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PROCEEDINGS

TECHNOLOGY TRANSFER CONFERENCE No.5

November 27 & 28 , 1984

Part 1
General Research

Organized By
The RESEARCH ADVISORY COMMITTEE

Sponsored By:

Policy and Planning Branch
A. Castel, Director

Organized By:

Research Advisory Coimnittee
P. D. Foley, Chairman

Program Coordinator:
Program Committee:

M. Moselhy

J. Bishop
D. Balsillie
M. Moselhy

Organization:

D. Bartkiw

D. Corr

M. Moselhy

J. Ralston

G. Schmidt

Administration:

B. Malcolm

Mailing Address:

POLICY AND PLANNING BRANCH
Research Coordination Office
135 St. Clair Avenue West
Toronto, Ontario

M4V 1P5

Telephone: (416) 965-5788

- 1 -

ABSTRACT
Recent Trends in Drinking Water Treatment

Vernon L. Snoeyink, Professor

of Environmental Engineering

University of Illinois

208 N. Romine
Urbana, Illinois 61801

Ontario Ministry of the Environment

Fifth Technology Transfer Conference

November 27 & 28, 1984

Holiday Inn-Toronto Airport

Toronto, Canada

The reaction to the finding that chlorination under certain
conditions resulted in excessive trihalomethane concentrations
has been a significant modification of chlorination practice.
Chlorine dose has been reduced, breakpoint chlorination has
been eliminated by some utilities, and many utilities now use
combined chlorine instead of free chlorine. These changes will
cause a greater frequency of problems caused by excessive
microbial growth in distribution systems, such as taste and
odor development, red water problems, and possibly regrowth of
bacteria that are measured by the coliform test. Such problems
will lead to a greater emphasis on water treatment to produce a
biologically stable water before it is distributed. Many
European utilities use biological water treatment processes
such as f luidized-bed and packed-bed fixed film reactors to
produce a water which can be distributed with low concentra-
tions of residual disinfectant-

A biologically stable water is one which does
not support growth of microorganisms to a
significant extent in treatment processes or
in the distribution system.

The main componants of biological instability
are NHJ , organic material, Fe^"*", and Mn^"*".

Biological instability can cause or enhance:

corrosion

growth of indicator organisms
taste, odor, and color production
reduced hydraulic capacity
disinfectant demand

Chlorination to control instability in
plant and distribution system

Produces THMs and TOX

Increases corrosion rates

Can cause chlorine taste and odor

High demand is costly and
accelerates other problems

- 5 -

£-Chlorophenol-Free Chlorine-GAC Reaction*

Compound

Identification
Method

OCH3

OOCH3

OH
CI

CI OH
OH CI

OH OH OH

(O>-0-^)-^ or (0>-0-(O>-&

CI CI CI C)

10.

- 6 -

(Continued)

OH OH OH

CI CI

OMe OH OH

CI CI CI

Identification
Compound ^ Method

^All compounds are identified in methylated samples.

A - Confirmed identification (Level 3) based on comparison of mass spectra
and retention times with those of authentic standards.

C - Confident identification (Level 2) based on comparison of mass spectra
with those of authentic standards from the literature.

D - Tentative identification (Level 1) based on mass spectra only; no
standards were available for comparison.

NH . -N Concentrations in Illinois
4

Surface Waters

Mean

0.21 mg/l

Median of Highest Location

0.86 mg/l

High Value

25.0 mg/l

Location with Median

Less than 0.25 mg/l

90%

Locations with Individual

Samples Greater than 1.0 mg/l

53%

NH^-N Concentrations in Illinois
Groundwater Supplies

Mean

0.62 mg/l

Mean of Highest County

3.2 mg/l

High Value

8.6 mg/l

Counties with mean

Less than 0.25 mg/l

39%

Counties with mean

Greater than 1.0 mg/l

21%

I
I

- 9 -

CHANGES IN CHLORINATTON PRACTICE

1. Reduce CI dose

2. Stop or modify prechlorination

3. Stop breakpoint chlorination

4. Add NH3, then CI

5. Add 01, then NH3

NH CL DISINFECTION EFFICIENCY

Cone. (mg/L)

time (99 % kill.min)

E. coli

Free CI

0.5

NH2CI

175
88

poliovirus

Free CI

JO.l

5
50

NH2CI

450
50

Ttfy\t^ of F/ocu

Croissy Biological Filter

Pozzolana media, 10-20mm/

ft ft ft t* !♦ 1^ ft ft

Effluent
to GAC
and
Disinfection

Air,

6.6 mVni^ Water

Aerated
influent

- 12 -

.5 m

4.75 m

6.1m

"lT

INLET

0.6 m

Performance

English

French

Biological

Pozzolana

Fluidized-Bed

Filter

Depth, m

Hydraulic load, m/d

108

240

Media diameter, mm

0.2

Influent NH4 -N, mg/l

3.2

Observed % Removals

T= 5°C

—

90-100

10°C

—

100

12°C

—

87-100

15°C

—

100

20°C

—

100

Summary

1. Achieving biologically stable water reduces chloriiiation
costs; production of THMs, TOX, tastes and odors;
regrowth; corrosion; and loss of hydraulic capacity.

2. Practical evidence and theoretical analysis shows that
biological process are efficient and reliable for removing
instability caused by NH4 .

3. Placing the biological process first is most advantageous,
becauce it allows easier operation of subsequent processes
and because subsequent processes provide multiple barriers
to microbial contamination.

- 15 -

DIRECT AND FOOD CHAIN UPTAKE OF ENVIRONMENTAL LEAD,
CADMIUM AND MERCURY IN A MODEL AQUATIC SYSTEM

Pamela Stolces
Institute for Environmental Studies, and
Department of Botany
University of Toronto

and

Peter Campbell

Instltut National de la Recherche Scientlflque

University of Quebec

Prepared for presentation to the Ontario Ministry of the
Environment, Research Advisory Committee, Technology Transfer
Conference No. 5, Toronto, November 27-28, 1984.

- 16 -

ABSTRACT

Recognising the theoretical and demonstrated relationships between
metal speclation and biological uptake on the one hand, and pH of water on
the other, we have reviewed the influence of pH on metal speciatlon In water
and the influence of pH on metal-surface interactions at the cell/medium
interface for a number of potentially toxic metals. Of these, sensitivity
of speclation to changes in pH should be low for Ag, Cd, Co, Mn, Ni and Zn,
moderate for Pb and high for Al, Cu and Hg; consideration of metal
dissolution or desorption from solid surfaces was not included in the
calculations. Supporting experimental evidence exists for Cu, Pb, Ag, Cd,
Mn and Zn. In terms of biological uptake, not only the chemical speciatlon
in solution but the effect of pH on the cell surface has to be taken into
account. From the limited amount of experimental data in the literature, it
can be shown that Cd, Cu and Zn are taken up (or exert toxic effects) more
at neutral pH than at acidic pH. The simplest hypothesis to explain this is
that competition occurs at the cell surface between the hydrogen Ion and the
metal ion. For Pb, low pH enhances uptake and toxicity.

Experiments on model food chains with defined media in which the
speciatlon can be calculated tend to support this hypothesis for the first
trophic levels (algae). It Is proposed that the observed relationship for
Hg, Cd and Pb In biota and pH in the field need to be considered
mechanistically in terms of hyrogen ion effects on metal speciatlon, metal
solubility, the cell surface, and (for all except algae) the indirect
pathway through the food chain.

1. INTRODUCTION

The acidification of aquatic and terrestrial ecosystems has a number of
subtle and potentially profound effects upon the biota of the affected
systems. As well as direct effects of the hydrogen Ion, changes in metal
concentration and availablility are known to be related to pH, and the
respective effects of hydrogen ion and metals can rarely be separated in
field studies.

Much progress has been made In studies of the chemistry of aquatic
systems over the past 10-15 years both in analytical techniques and the
modelling of thermodynamics and kinetics of chemical reactions in solution
(Stumm and Morgan, 1981). The significance of this type of information to
biological studies related to nutrients of toxic substances means that it is
now axiomatic that a 'total* value for an element in water is rarely
adequate to predict the potential biological activity of that element.
While the modelling approach is useful to this end for defined systems, we
are still left with a sense of inadequacy In terms of assessing biological
activity of elements in an undefined matrix such as natural water. The
bioassay or the biomonitorlng approach Is still the ultimate test of
toxicity and availability.

An attempt was made recently to Identify metals of potential or
demonstrated effects in relation to acidification (Campbell e£al., 1983)
and was referred to briefly at this meeting last year (Stokes and Richman,
1983). Since then we have attempted to refine our approach to the

- 17 -

relationship between pH and metals, and the present paper concentrates on
theoretical and experimental work rather than field studies. Field studies
have shown repeatedly that there is often a negative corraelatlon between
water pH and Hg in fish (Suns et^ al . , 1980; Hakanson, 1980; Wren and
MacCrimmon, 1983), algae (Stokes e£al., 1983) and invertebrates (Hultberg,
personal communication) and there are indications that cadmium and lead
follow similar patterns for fish (Suns, personal communication). In one
study, lead showed the same pattern in algae (Stokes e£ al., 1983), even
though total lead in water was not related to pH in the algal study (Bailey
and Stokes, 1984). The biota in the field integrate over time the combined
effects of the hydrogen ion on water chemistry and on the organisms' uptake
processes. The following study emphasises an attempt to sort out these two
classes of components, each of which is clearly very complex. In order to
simplify this, we have concentrated on the changes in chemical speciation
rather than changes in total concentration (i.e. geochemlcal mobilisation)
as affected by pH.

The approach will be presented firstly from the theoretical point of
view, secondly from a selection of examples from the literature, and finally
an experimental approach in our own laboratories will be described.

2. THEORETICAL CONSIDERATION

Even at constant metal concentration, changes in metal speciation can
be anticipated as a result of decrease in pH by a shift in hydrolysis
equilibria favouring the aquo-ion, shift in complexation equilibria, and
shift in specific adsorption equilibria.

Using theromdynamlc calculations (e.g. Jenne, 1979), based on
•synthetic* lake water, defined and resembling softwater precambrian shield
lake composition (Campbell et^ al . , 1983) the MINEQL-1 chemical equilibrium
model (Westall et^ al^. , 1976) was used to calculate the theoretical
speciation of Ag, Al, Cd, Co. Cu. Hg, Mn, Nl, Pb and Zn (Campbell et al.,
1983). Simulations were performed for aerobic conditions, at fixed pH
values of 4, 5, 6, and 7, with no organic ligands and no adsorbing surfaces.
Table 1 shows some of the results. In summary, pH-related changes in
speciation are expected for Al, Cu, Hg and Pb, but not for Ag, Cd, Co, Mn,
Nl and Zn.

Concerning the cell surface, the functional groups present which would
bind metal ions (Crist e£ a]^. , 1981) would also provide sites for binding
hydrogen ions, with the result that the protons could play a role analagous
to that of calcium ions in protecting the cell against toxic metal uptake
(Pagenkopf, 1983). This is considered In greater detail elsewhere (Campbell
and Stokes, in preparation) but in effect what it means is that while the
hydrogen ion can increase the availability of certain metals, it may also
decrease the actual uptake of the metal. An exception to this would be
mercury when in the methyl form, since this moves into the cell in the
organic form and is probably uncharged.

3. EXAMPLES FRCM THE LITERATURE

We have attempted to locate examples of studies relating to the three

- 18 -

metals identified as major concerns in fish but also considered examples
based on the speciation behaviour from table 1. Cadmium, it should be
noted, is not expected to change its speciation over the pH 7-4 range.
Several examples of studies with algae show that the lower the pH, the less
the cadmium uptake (Table 2). This therefore supports the hypothesis
concerning competition between metal Ions and hydrogen ion. Interestingly
it appears to apply also to fish (Table 2), In short term bloassays where
presumably direct uptake is more significant than food chain uptake.

For Pb the data tend to show that lower pH enhances uptake and toxicity
for algae (Monahan, 1976), fish (Merlin! and Pozzl , 1977) and fungi (Bablch
and Stotsky, 1979). Lead, unlike cadmium, is expected to be more available
at low pH, and the solubility of lead is also strongly pH dependent. We
have also reviewed the more extensive literature on Zn (which does not
change its speciation) and Cu (which does) and for each of these there Is a
great deal of evidence to support the idea of a competitive mechanism such
that uptake and/or toxicity to algae, to fish and in a few examples to
bacteria, are greater at neutral than at low pH (Stokes and Campbell, 1984).

4. LABORATORY STUDIES

Our experimental design was described earlier (Stokes and Rlchman,
1983); briefly we are measuring uptake (direct) from synthetic lake water
(based on the chemical composition used in the MINEQL model calculation) of
Cd (0.05 mg L-1), Pb (0.10 mg L-1) and Hg (0.05 and 0.10 ug L-1), I.e.
realistic levels, into algae ( Oocystls marsonll ), amphipods ( Hyallela
azteca ) and fish ( Perca flavescens ) at pH 5-5 and 7.0 respectively. We also
look at the potential for metal uptake via food for the two consumer levels,
again at two pHs .

To date the results indicate that uptake into algae is rapid , and for
Cd Is slightly higher at pH 7.0 than at 5.5. For lead we found no effect of
pH. The bioconcentration factors (OF) for Pb and Cd are of the order of
100, which is less than we found for lead in field studies, but those of
course included sediment which probably provided a source of metal
especially in acidified systems (Stokes et_ al., 1983). For methyl mercury,
the CFs were much greater (10, OOOx) but uptake did not appear to be pH
dependent over the 5-7 range.

Values for amphipods and fish have not yet been determined, so no
conclusions can be drawn for these trophic levels at this time.

5. CONCLUSIONS

The results of laboratory experiments in simplified model systems
indicate that for Cd, Zn and Cu, direct uptake of metal by algae and for a
few examples by fish is affected by pH such that lower pH tends to decrease
uptake, even for copper whose speciation is pH sensitive with lower pH
increasing bioavailability. This is at odds with the observations for fish
in acidic lakes, in which the Cd is negatively correlated with pH. For
mercury, laboratory experiments have not demonstrated any clear difference
in uptake between pH 5 and 7. For lead, the experimental data agree with
the field data; the lower pH enhances uptake and toxicity.

- 19 -

Field values for body burdens of metals incorporate the influence of pH
on metal mobility (solubility), metal speciation and the interaction of
hydrogen ion with metal ions at the cell surface, and, for consumers, the
effect of pH on the indirect route of metal uptake via food. It is also
likely that physiological processes are affected over the range 7-4. The
result of the direct and food uptake experiments at different pHs in the
present study combined with the direct uptake data already available may
shed some light on the mechanisms Involved.

ACKNOWLEDGEMENTS

Ue wish to thank Gllles Groulx, Lisa Richman, Mel Martin and Kit Yung
for technical assistance and Penny Ashcroft Moore for typing the manuscript.
The practical work is being supported by the Ontario Ministry of the

Environment.

6. REFERENCES

Babich, H. and G. Stotsky. 1977. Abiotic factors affecting the toxicity of
lead to fungi. App. Env. Microbiol. 506-513.

Bailey, R.C. and P.M. Stokes. 1984. Evaluation of filamentous algae as
biomonitors of metal accumulation in softwater lakes: Multivariate
approach. Aquatic Toxicilogy and Hazard Assessment: Seventh Symposium.
ASTM STP 854, R.D. Card we 11 , R. Purdy and R.C. Bahner (eds.) American
Society for Testing and Materials, 1984 (in press).

Campbell, P.G.C., P.M. Stokes and J.G. Galloway. 1983. Effects of

atmospheric deposition on the geochemical cycling and biological
availability of metals. Proc . Inter. Conf. on Heavy Metals in the
Environment. September 6-9, 1983.

Crist, R.H., K. Aberholser, N. Shank and M. Nguyen, 1981. Nature of bonding
between metallic ions and algal cell walls. Environ. Sci. Technol.
J^(IO): 1212-1217.

Cusimano, R.F., D. Brakke and G.A. Chapman. 1984. Unpublished manuscript.

Gipps, J.F. and B.A.W. Coller. 1980. Effect of physical and culture

conditions on uptake of cadmium by Chlorella pyrenoidosa . Aus. J- Mar.
Freshwater Res. 3]^; 747-755.

Hakanson, L. 1980. The quantitative impact of pH, bioproduction and Hg-
content of fish (pike). Environ. Pollut . (Ser. B) U 285-304.

Jackson, T.A., G. Kipphut, R.H. Hesslein and D.W. Schindler. 1980.

Experimental study of trace metal chemistry in softwater lakes at
different pH levels. Can. J. Fish Aquat. Sci. ^: 387-402,

Jenne, E.A. 1979. Chemical modelling in aqueous systems. A.C.S. Symp. Ser.
93, Washington, D.C. American Chemical Society.

- 20 -

Les, A. and R.W, Walker. 1984. Toxicity and binding of copper, zinc and

cadmium by the blue green alga Chrococcus paris . Water, Air and Soil
Pollut. ^: 129-139.

Monahan, T.J. 1976. Lead inhibition of chlorophycean mlcroalgae . J.
Phycol. I2i 358-362.

Merllni , M. and G. Pozzi. 1977. Lead and freshwater fishes. Part I: Lead
accumulation and water pH. Environ. Pollut. 12: 167-172.

Pagenkopf, G.K. 1983. Gill surface interaction model for trace-metal
toxicity to fishes. Role of complexation, pH and water hardness.
Environ. Sci . Technol. J7; 342-347.

Peterson, H.G., F.P. Healey and R. Wagemann, 1984. Metal toxicity to algae
a highly pH dependant phenomenon. Can, J. Fish. Aquat . Sci- 41(6):
974-979.

Sakaguchi, T., TA. Tsuji, A. Nakajima and T. Horikoshi . 1979. Accumulation
of cadmium by green mlcroalgae. European J. Appl . Microbiol.
Biotechnol. 8^: 207-215.

Stokes, P.M., R.C. Bailey and G.R. Groulx. 1983. Metals in acid-stressed
and other softwater lakes, with an evaluation of attached filamentous
algae as biomonitors. Report to OMOE, March, 1983. 48p.

Stokes, P.M. and P.G.C. Campbell. 1984. Acidification and toxicity of

metals to aquatic biota. Presented at 11th Aquatic Toxicity Workshop,
Vancouver, B.C., November, 1984.

Stokes, P.M. and L. Richman. 1983. Partitioning of mercury, lead and

cadmium in aquatic systems, in relation to acidification, ^n Proc.
OMOE Technology Transfer Conference, Constellation Hotel, November,
1983.

Stumm, W. and J.J. Morgan. 1981. Aquatic Chemistry. John Wiley and Sons.
780 p.

Suns, K. , C. Curry and D. Russel . 1980. The effects of water quality and
morphometric parameters on mercury uptake by yearling perch. Ontario
Ministry of the Environment Technical Report LTS 80-1.

Westall, L.M., J.L. Zachary and F.M.M. Morel. 1976. MINEQL, a computer

program for the calculation of the chemical equilibrium composition of
aqueous systems. M.l.T. Civil Eng, Tech. Report 18 : 91p.

Wren, CD. and H.R. MacCrimmon. 1983. Mercury levels in sunfish ( Lepomls
gibbosus ) relative to pH and other environmental variables of
Precambrian Shield lakes. Can. J. Fish. Aquat. Sci. 40: 1737-1744.

- 21 -

TABLE 1

Calculated change In speciatlon in a low conductivity
soft water inorganic medium, no sediments , pH 7-A

Concentration* in
Metal model x KT^M

0.10

58.90

1.00

Co(II)

0.10

1.00

0-01

7.24

1.00

0.05

5.00

Predicted sensitivity
to pH change

Observed**
changes

low

high

low

high

low

moderate

high

low

* Based on realistic concentrations in natural waters.
** Jackson et^ al , 1980.
? no clear trend.

- 22 -

TABLE 2

Effect of. £H on cadmium uptake and toxicity:
selected laboratory studies

Organism

Reference

Results

Chlorella

pyreaoldosa 8.3.7.3,6.6 Gipps and Coller , 1980

Chlorella
regularls

Scenedesmus
quadlcauda

Salmo

Chroococous
parls

7.3

Sakaguchl et al . , 1979

8.5-5.5 Petersen et al., 1984

galrdnerl 7,5.7,4.7 Cuislmano et al., 1984

7,6.5,4 Les and Walker, 1984

Cd uptake and toxicity
least at 6.6, greatest
at 7.7

More Cd uptake at 7 than
at 3

More toxic at high than
low pH: 200-fold
increase over range
5.5-8.5.

tC=,Q at 7«<0.05 at
4.7-2.8 ug L-1 Cd.

Uptake decreased as pU
decreased.

- 23 -

STUDIES OF THE NITRATE DISTRIBUTION

AND NITROGEN TRANSFORMATION

IN A SHALLOW SANDY AQUIFER

NEAR ALLISTON. ONTARIO

R.W. Gillham, R.C. Starr, F.F. Akindunni and S.F. O'Hannesin

Institute for Groundwater Research

Department of Earth Sciences

University of Waterloo

Waterloo, Ontario

N2L 3G1

- 24 -

Acknowledgements
Wg were first introduced to the potential contamination problem in
the unconfined aquifer at Alliston by Dr. Alan Hill of the Geography
Department at York University. His cooperation has been most helpful.
Jim Barker and Colin Mayfield, of the Earth Sciences and Biology Depart-
ments, respectively. University of Waterloo, have provided expert advice
and assistance in the design of experiments related to an evaluation of
factors affecting the deni trification process. Without the cooperation
of the private land owners, this study could not proceed. We are par-
ticularly grateful for their tolerance and assistance.

- ii -

- 25 -

STUDIES OF THE NITRATE DISTRIBUTION AND NITROGEN TRANSFORMATIONS
IN A SHALLOW SANDY AQUIFER NEAR ALLISTON. ONTARIO

R. Gin ham , F. Akindunni, R. Starr, S. O'Hannesin
TnFtitute for Groundwater Research
Department of Earth Sciences
University of Waterloo
ABSTRACT Waterloo, Ontario N2L 3G1

A recent survey showed a high proportion of domestic groundwater supplies in the
unconfined sand-plain aquifer near Alliston to contain nitrate concentrations in ex-
cess of the drinking water limit of 10 mg/L NO3 -N. This study was undertaken to
provide a more complete description of the spatial distribution of nitrate in the
aquifer, to examine the factors responsible for the highly variable concentration
distribution and to develop methods of groundwater development that would result in
domestic supplies of Improved quality.

Monitoring devices installed at several locations in the watershed showed the
nitrate concentration versus depth to be highly variable. Under shallow water-table
conditions, nitrate contamination was generally limited to depths of two to four
meters below the water table. The decline in the nitrate concentration was generally
matched by a decline in the dissolved oxygen concentration. These trends were simi-
lar to those observed elsewhere, and suggest that the depth of penetration of the
contaminated zone is limited by the denitrification process. In areas with deep
water tables, nitrate and significant concentrations of dissolved oxygen occurred
across the entire saturated thickness of the aquifer. Laboratory tests conducted
on core samples of the aquifer material suggested that under shallow water table
conditions, sufficient labile organic carbon was transported to the water table to
cause a substantial reduction in the dissolved oxygen concentration. With the
development of reducing conditions, nitrate was converted to nitrogen gas by
denitrification. Under deep water-table conditions, the data suggests that there
was insufficient labile organic carbon reaching the water table to result In the
development of reducing conditions and thus denitrification would not be a signi-
ficant process in these environments.

Field tests indicated that in situations with stratified nitrate contamination,
a substantial Improvement in the quality of domestic supplies could be achieved
simply by Installing the intake zone of the wells at greater depth in the aquifer.

- 26 -

Introduction

Nitrate is the most widespread of the recognized groundwater con-
taminants. Potential sources of contamination include point sources
such as septic fields, livestock feedlots and waste lagoons, while agri-
cultural fertilizer is undoubtedly the most significant of the distrib-
uted sources. Point sources generally result in localized zones of con-
tamination, and although these can have a very detrimental effect on
local water supplies, remedial measures such as removal of the source,
or changing the location of the well are frequently possible. On the
other hand, distributed sources generally result in extensive zones of
contamination, significantly reducing the possibility for effective rem-
edial measures.

During the period from 1976 to 1980, the Department of Earth Sci-
ences at the University of Waterloo conducted extensive surveys of the
nitrate distribution in groundwater associated with agricultural wat-
ersheds and undertook more detailed studies concerning the chemical
transformations of nitrogen in hydrogeologic regimes. The results of
those investigations are included in Hendry et al . (1932), Trudell
(1980), Gillham and Cherry (1978) and Egboka (1978). In summary, it was
shown that extensive nitrate contamination occurs in areas with perme-
able soil materials, that are under intensive agricultural production.
Areas showing significant nitrate contamination on a regional scale
included sandy, unconfined aquifers near Leamington, Harrow, Delhi and
Rodney. The detailed studies have shown that where the water table

- 27 -

occurs at depths of less than about 3 or 4 m below ground surface,
nitrate contamination is limited to a depth of 2 to 3 m below the water
table. Generally a sharp decrease in nitrate concentration, dissolved
oxygen and redox potential occurs at a relatively shallow depth below
the water table. This evidence, plus detailed tracer tests and geochem-
ical studies conducted by Trudell (1979) shows the loss of nitrate with
depth to be the result of denitrification; the biochemical conversion of
nitrate (NOI) to nitrogen gas (N^). with the conversion to nitrous oxide
(N^O) as an intermediate step. In areas with deeper water tables, there
appears to be a gradual decrease in nitrate concentration with depth;
however, the trends are not as distinct as in areas with shallow water
tables, suggesting that denitrification is not as active in these
regions. Though specific tests have not been conducted, we believe that
the reduced rates of denitrification under deep water-table conditions
may be related to a reduced availability of organic carbon.

The biochemical conversion of nitrate to nitrogen gas requires that
the appropriate bacteria be present, that reducing conditions prevail
and that an organic energy source be available for bacterial respira-
tion. An understanding of the occurrence of denitrification therefore
requires an understanding of the processes and factors that contribute
to the occurrence of these conditions in natural groundwater regimes.

In a hydrogeologic context, denitrification is highly beneficial in
that it converts nitrate, which Is toxic to Infants and livestock
(drinking-water standard of 10 mg/L NO! - N) to nitrogen gas which is
not toxic. An understanding of the processes that cause denitrification
to occur in natural hydrogeologic environments could therefore make a

- 28 -

substantial contribution to an improvement in the quality of water sup-
plies situated in nitrate-sensitive areas. In particular, it could
resul t in al tered or improved land-management practices that would
reduce the amount of nitrate entering the groundwater system, or could
contribute to the development of improved methods of groundwater devel-
opment in these areas.

A recent survey of domestic wells conducted by Dr. A.R. Hill of
York University has revealed extensive nitrate contamination in a sandy
aquifer near Alliston, Ontario. Of 164 groundwater samples collected,
68 exceeded the drinking-water standard of 10 mg/L, with concentrations
as high as 95 mg/L being recorded. The results showed that 40% of the
aquifer was contaminated, and that the contaminated area was highly cor-
related with the production of potatoes and other high-value agricultur-
al crops. Because domestic wells were used in the survey, there was
1 ittle information concerning the variation in nitrate concentration
with depth below the water table; however, there was some indication of
a decreasing concentration with depth. This would be consistent with
the results obtained in other areas by the University of Waterloo.

The shallow aquifer at Alliston is used extensively for domestic
water supplies. The results of the survey conducted by Dr. Hill there-
fore reveals an extensive and serious incidence of groundwater contami-
nation. If denitrification is an active process in the aquifer, then it
is quite possible that the degree of contamination and the extent of the
contaminated zone will not increase in the future. In this event, rem-
edial measures may be possible within the existing conditions. On the
other hand, if denitrification is not a significant process, and assum-

- 29 -

ing that the land- use practices do not change substantially, it is
probable that in time, the entire aquifer will become contaminated. In
this event, the residents will require an alternate water supply.

The present study was undertaken to examine the distribution of
nitrate in the surficial aquifer at All is ton, and to investigate the
occurrence of denitrification in the aquifer.

During the first year of the study, four activities were initiated:

1. The first activity was undertaken to obtain an improved definition
of the spatial distribution of nitrate at selected locations in
the aquifer.

2. Where denitrification is a significant process, nitrate contamina-
tion appears to be limited to the upper few meters of the saturat-
ed zone. This activity investigated methods of groundwater devel-
opment that would yield domestic supplies of acceptable quality in
aquifers of stratified contamination.

3. The second activity was designed to identify and investigate those
factors that appear to control the occurrence of denitrification
in unconfined aquifers,

4. As a possible aid to activity 3, a device was developed for the in
situ measurement of denitrification rates in saturated geologic
materials.

In this paper, the progress during the first year of the investiga-
tion concerning activity 1 and 2 will be summarized very briefly, while
the results pertinent to activity 3 will be presented in somewhat great-
er detail. Further details of all the activities are presented in the
1982-83 progress report submitted to the Ontario Ministry of the Envi-
ronment (Gillham et al., 1984).

- 30 -

Nitrate Distribution in the Unconfined Alliston Aquifer
Three study areas, number 1. 2 and 3 in Fig. 1, were selected In
the Alliston sand plain. Based on the background hydrogeologic Informa-
tion that was available, it was expected that the geochemical conditions
and nitrate distributions at area 1 would be representative of shallow
water-table conditions, those at site 2 would be representative of deep
water-table conditions, while area 3 would contain regions that would be
representative of both deep and shallow conditions. Seven sampling
wells, each containing about 8 sampling points spaced at vertical inter-
vals of about 50 cm were installed at area 1. three were installed at
area 2 and 6 at area 3. This gave a total of approximately 130 sampling
points. Samples were collected from all sampling points on several
occasions and analysed for NO^ and CI". Samples from selected wells
were analysed for dissolved oxygen and tritium.

The most striking characteristic of the data was the great vari-
ability in the nitrate concentration with depth. Generally the maximum
concentrations occurred near the water table {greater than lOOmg/1 NO3-N
in some cases) then declined, frequently to values approaching zero, at
greater depth. Even areas where the evidence suggested that denitrifi-
cation is not an active process, or is proceeding at a slow rate, showed
substantial declines in nitrate concentration with depth. It is there-
fore apparent that water-supply wells are an inappropriate source of
samples for characterizing nitrate distributions in contaminated aqui-
fers. A screen located at depth in the aquifer would not detect the

- 31 -

presence of nitrate, a screen located near the water table would show
maximum levels of contamination, while wells screened over the entire
thickness would give some average value. In many cases the average val-
ue could be less than the drinking water limit (10 mg/1 NOl-N) even
though zones contaminated to levels well in excess of the limit may
exist.

All the areas studied to this point show some degree of nitrate
contamination. It is reasonable to expect this to be the situation
throughout the cultivated areas of the Alliston sand plain. The extent
and severity of contamination appears to be the greatest in areas with
"deep" water tables where denitrlfication is not an active process, and
somewhat less in areas of shallow water tables.

As a result of denitrlfication, at the sites having shallow water
tables, nitrate contamination is generally limited to shallow depths
below the water table, thus providing significant protection for the
quality of the deeper groundwater. Where denitrlfication does not
appear to be a significant process, nitrate has penetrated to considera-
bly greater depths in the aquifer. The data suggests the "shallow" con-
dition to be associated with water tables that are less than about 3 to
4 m below ground surface.

- 32 -

Alternate Methods of Aquifer Development
Because of the stratified nature of the contamination, it was pos-
tulated that in many instances installation of the screened interval of
the well at greater depth would result in an improvement in the quality
of domestic supplies. For this study, a site was selected in study area
1 that clearly showed stratified contamination, with high nitrate con-
centration at shallow depths below the water table, and low values at
greater depth. Three wells with 1-m long screens were installed, one
having the screened interval in the contaminated zone, one with the
screened interval spanning the interface between the contaminated and
uncontaminated zones and the third was positioned with the screen in the
uncontaminated zone. A fourth well was installed with a long screen
that extended over almost the entire saturated thickness of the aquifer.
Each well was pumped at different rates and for different durations and
the concentration of nitrate in the discharge was recorded as a function
of time. In general, the results supported the suggestion that the
qual i ty of domesti c suppl i es coul d be improved by i nstal 1 i ng the
screened interval of the wells at greater depth.

Conclusions based on the results of these tests must be qualified
in that the pumping schedule was very different from that of a domestic
well. Generally, a domestic well would be pumped for a relatively short
period of time, then would rest (recover) for a period of time. The
length of the pump and rest phases could be highly variable, depending
upon the users needs and would vary both diurnally and seasonally. No

- 33 -

attempt was made to simulate this type of use, though it was felt that
at the flow rates used, a continuous pumping period of six hours as used
in the tests would represent a relatively severe test.

To further quantify the effects of screen placement on the quality
of the discharge water and to examine the sensitivity to factors such as
depth below the contaminated zone, screen length and pumping character-
istics, simulation models may provide the most practical investigative
procedure. Nvertheless, based on the preliminary study, we suggest that
drive points should not be used for obtaining domestic water supplies,
and that many of the existing wells should be deepened.

- 34 -

Investigation of Factors Controlling the Occurrence of

Denitrification
Introduction

As noted previously, denitrification appears to be a significant
process in unconfined aquifers with shallow water tables, but is a rela-
tively minor process under deep water-table conditions. We suspect that
under shallow water-table conditions, substantial organic carbon is
leached to the water table where it is oxidized, giving rise to reducing
conditions and thus an environment in which denitrification can proceed.
Under deep water-table conditions, it is suggested that much of the
labile organic carbon is oxidized prior to reaching the water table and
thus the reducing conditions necessary for denitrification are not pro-
duced.

In order to investigate the above hypothesis, soil samples from
different depths in the vadose zones of a shallow water-table site and a
deep water-table site were collected and the ability of these samples to
support denitrification in the laboratory was determined. The prelimi-
nary results obtained to date are consistent with the hypothesis stated
above.

Theory

Denitrification is the microbiological process by which nitrate is
converted into molecular nitrogen gas. Chemoheterotrophic bacteria per-
form this process by using nitrate to oxidize organic carbon. If dis-

- 35 -

solved oxygen is present in appreciable quantities, the microorganisms
use oxygen in preference to nitrate as an oxidizing agent; therefore,
denltrification occurs only if the dissolved oxygen concentration is

low.

Denitrifying microorganisms utilize an organic carbon substrate as
a source of carbon for the production of biomass and as a source of
energy. Some of the organic carbon is converted to carbon dioxide, so
the organic carbon concentration decreases with time unless additional
organic carbon Is added. The major source of organic carbon in natural
soils is decaying plant matter, which is more abundant at shallower
depths. Microorganisms utilize the more labile fraction of the degrada-
tion products in preference to the less labile fraction, so the organic
carbon pool In a closed system becomes more refractory with time.

Water in the vadose zone generally flows vertically downward, so
that water at depth has a greater residence time 1n the vadose zone than
water at shallower depths. Infiltrating water dissolves organic carbon
In the top few tens of centimeters of the profile where there is abun-
dant plant matter. Little carbon is added at greater depth, so a slug
of water in the vadose zone below a depth of a few tens of centimeters
can be regarded as a closed system with respect to the input of addi-
tional organic carbon. The dissolved carbon is utilized as the water
moves downward, so the organic carbon fraction becomes more refractory
and total concentrations of dissolved organic carbon decrease with depth
In the vadose zone.

Based on our previous studies, it Is suggested that the trend
towards lower concentrations of DOC and a more refractory character at

- 36 -

depth is the mechanism that controls the denitrification rate in aqui-
fers. The organic carbon transported into an aquifer overlain by a thin
vadose zone will be more labile and at higher concentrations than the
organic carbon transported into an aquifer overlain by a thicker vadose
zone. The higher concentration of more labile organic carbon is able to
support a more active population of microorganisms that first consume
the available oxygen and then turn to nitrate as an oxidizing agent.
With a thicker vadose zone and its lower concentrations of labile organ-
ic carbon, the microbial population would be less active. If the vadose
zone was thick enough so that the concentration of labile organic carbon
entering the underlying aquifer was so low that the labile carbon was
consumed before all of the dissolved oxygen was utilized then denitrifi-
cation would not occur in the aquifer.

Procedure

Two field sites were used in this phase of the investigation. Both
are active agricultural areas underlain by glaciofluvial , medium-to-
fine-grained sand. The site near Rodney, Ontario has a vadose zone 1
meter thick, and the site near Alliston has a 4.5 meter thick vadose
zone. It would be expected that more labile carbon would enter the
aquifer at Rodney than at Alliston. and therefore that denitrification
would proceed at a greater rate at Rodney,

Soil samples were collected from the vadose zone of each site and
returned to the laboratory for evaluation of their ability to support
denitrification. This was determined by placing soil in a flask and
adding water, nitrate and nutrients other than carbon. The flask was

- 37 -

sealed and an oxygen-free atmosphere was created in the headspace. If
sufficient labile organic carbon was present in the flask, the dissolved
oxygen was consumed and denitrifi cation proceeded. The amount of
nitrate consumed is therefore a quantitative index of the ability of a
particular soil sample to support deni tri fication. If the hypothesis is
correct, shallow soil samples would support a large amount of denitrifi-
cation and a large amount of nitrate would be utilized. Deeper soil
samples could support less denitrification and therefore less nitrate
would be utilized.

The acetylene inhibition method was used to quantify the amount of
nitrate consumed. The pathway by which nitrate is converted to nitrogen
is

NO3 — NO2 — N^O — N2
If acetylene is present in the system, the conversion of nitrous oxide
to nitrogen gas (Np) is blocked. Assuming that the NO^ -- NO2 -- N^O
pathway is the only significant source of nitrous oxide in the flasks,
then the N^O produced is proportional to the NO3 utilized by the deni-

trifiers.

Acetylene was added to the flasks in the laboratory. Samples of
the headspace gas were collected after about 40, 100 and 200 hours of
incubation and analysed for N2O by gas chromatography.

Similar flasks were set up with a readily available carbon source,
glucose, added to selected soil samples. This was done in order to dem-
onstrate that carbon rather than the microbial population was the limit-
ing factor. Control flasks were also included; these flasks had neither
nitrate nor carbon added.

- 38 -

Results and Discussion

Nitrate and Dissolved Oxygen Profiles

Concentration profiles below the watertable suggest that denitrifi-
cation is actively occurring in the aquifer near Rodney, but either not
at all or only at depth in the aquifer at Alliston. The concentration
of nitrate and dissolved oxygen at Rodney, Figure 2, shows that only
trace amounts of nitrate are present a few centimeters below the water-
table, and non-detectable concentrations below about 1.4 m. The concen-
tration of dissolved oxygen decreases from near saturation to less than
2 mg/1 over this interval. Considering the probable high nitrate load-
ing due to fertilizer application, the low nitrate concentration and the
decrease in dissolved oxygen strongly suggest that deni tri fication is an
important process at this location. In contrast. Figure 3 shows that at
the Alliston site, nitrate and dissolved oxygen concentrations persist
at high concentrations for 8-10 m below the water table, suggesting that
denitrification is not an important process at shallow depths in the
aquifer.

Incubation Tests - Rodney (Shallow Water Table)

Figure 4 shows typical results of the incubation of soil samples
using the acetylene inhibition method. Results from only two depths,
.02 and .77 m are included. The plot labeled 'NO3' represents samples
to which only nitrate was added, while both nitrate and organic carbon
were added to the samples represented by the plot labeled 'C NO3'. For
the sample collected at shallow depth, 0.02 m (Fig. 4a), and to which

- 39 -

both carbon and nitrate were added, the concentration of N^O approached
a maximum value within the first 40 hours then remained relatively con-
stant for the duration of the test. This indicates that the initial
rate of denltrification was ^ery high, followed by a rate of near zero
after a time of 40 hrs. The initially high rate indicates that neither
the bacterial population or the available carbon were rate limiting,
while after about 40 hrs, some nutrient, probably nitrate, was limiting.
For the sample to which only nitrate was added, the rate of denltrifica-
tion was reasonably constant over the first 100 hrs, then decreased.
Comparing the two curves of Fig. 4a suggests that for the shallow soil
sample, available carbon presents a slight limitation to the rate of
denltrification. Nevertheless, even in the sample to which no carbon
was added, the rate of denltrification was high.

For the sample collected at a depth of 0,77 m (Fig. 4b), the ini-
tial rate of denltrification was low for both samples. For the sample
to which both carbon and nitrate were added, the rate increased rapidly,
while for the sample to which only nitrate was added, the rate remained
low and at a relatively constant value over the 200-hr duration of the
test. The result suggests that the rate of denltrification is limited
by the available carbon, and further, that at early time, the rate may
also be limited by the microbial population.

Figure 5a shows the N«0 produced after 104 hrs of incubation versus
depth. As suggested by Fig. 4, at shallow depth, neither the available
carbon nor the microbial population were a limitation to N^O production
(denltrification). At greater depth, down to about 50 cm, it appears
that carbon is the limiting factor, while at yet greater depths, both

- 40 -

the available carbon and the microbial population are limiting factors.
It should be noted that had the data at a time of 200 min. been plotted
in Fig. 5, the N«0 produced in C + NO^ samples would have been almost
constant with depth, giving no indication of limitations as a result of
the bacterial population. At the time scale of field processes, it
appears reasonable to conclude that over the depth interval considered,
the presence or absence of suitable bacterial populations would not be a
limitation to the denitrification process. In addition, even though the
rate of denitrification is low at depth and is apparently limited by the
availability of a suitable carbon substrate, the rate is nevertheless
significant and could account for a substantial loss of nitrate over a
period of weeks or months.

Figure 5 shows the TOC concentration in pore water extracted from
the soil samples versus depth. With the exception of the sample col-
lected at the shallowest depth, the decline in TOC parallels the decline
in N^O production in the samples to which only nitrate was added.
Though far from conclusive, this suggests a relationship between rates
of denitrification and TOC concentration and is consistent with our ear-
lier hypothesis that the occurrence of denitrification is related in
some manner to the mechanisms responsible for the translocation of

organic carbon.

Figure 6a is a typical N^O-depth profile determined from the incu-
bation procedure, using core material collected at the Alliston site.
The results are in general agreement with the results from Rodney, with
high denitrification rates at shallow depth and lower rates at depth.
The graph of TOC versus depth (Fig. 6b) suggests that denitrification is

- 41 -

carbon limited. In contrast to Rodney, the samples from the lower por-
tion of the vadose zone, below 3,4 m, had non-detectable denitrifica-
tion rates, suggesting that the rate of denitri fication in the underly-
ing aquifer would be very slow.

Anomalously high rates were observed from 2.4 to 3.1 m. It is
believed that this anomaly was not caused by sample contamination, but
is due to high carbon centrations at this depth in the vadose zone, as
shown in Figure 6b. The origin of this high carbon content is presently
unknown.

Summary of Results

The results of preliminary analyses of samples from the vadose zone
at shallow and deep water-table sites at Rodney and AUiston support the
hypothesis that the depth of the watertable affects the rate of denitri-
fication in groundwater. The rate-controlling mechanism appears to be
the availability of labile organic carbon in the aquifer. The input of
labile organic carbon depends on the vadose zone thickness; as the
vadose zone thickness increases less labile organic carbon enters the
underlying aquifer. Therefore in unconfined aquifers with a shallow
water table, sufficient labile organic carbon enters the aquifer to sup-
port significant rates of denitrification. With a deeper water table,
the amount of labile organic carbon entering the groundwater is able to
support only much lower rates of denitrification.

- 42 -

References

Egboka, S.. 1978. Field Investigations of deni trification in groundwa-
ter. M.Sc, thesis. Department of Earth Sciences, University of
Waterloo, 149 pp.

Gillham, R.W,, and J,A. Cherry, 1978. Field evidence of denitrification
in shallow groundwater flow systems. In: Proceedings of the Thir-
teenth Canadian Symposium on Water Pollution Research in Canada, v.

13, p. 53-71.

Hendry. M.J., R.W. Gillham and J. A. Cherry, 1983. An integrated

approach to hydrogeologic investigations: A case history. Journal
of Hydrology, v. 63, pp. 211-232.

Trudell, M.R., 1980. Factors affecting the occurrence and rate of deni-
trification in shallow groundwater flow systems. M.Sc, thesis.
Department of Earth Sciences, University of Waterloo.

- 43 -

Figure 1. Locations of study sites in the Alliston aquifer,

- 44 -

M03 LM&/L fl5 NJ

o
o

_ _^

00 LM&/LJ
J). 00 S. 00

Figure 2. Nitrate and dissolved oxygen profiles at the Rodney site.

- 45 -

N03 (HG/L flS N)
cP. 00 50.00 100.00

— I 1 ( 1

■i I

0.00

Figure 3. Nitrate and dissolved oxygen profiles at the Alliston site.

- 46 -

0-00

!iO. 00 lOaOO 16U 00

riM£ [HDUBSJ

0. 00

^C. UO iOO. 00 IS'C. 00

TiME IHOUftSJ

200. 00

Figure 4. Nitrous oxide production versus time in the incubation tests using
samples from the Rodney site, a) 0.02 m depth, b) 0.77 m depth.

- 47 -

n:o

,0.00

tPPM) LXI03 J

10. OC 2D, 00

TOC

,0.00

CMG/LJ

50.00

00.00

Figure 5. Rate of production of nitrous oxide, and dissolved organic carbon
concentration versus depth at the Rodney site.

- 48 -

,0.00

roc

LM&/LJ
5"0.00

rs>

m s

o
o

C + NO.

* NO.

o
o

100.

Figure 6. Rate of production of nitrous oxide and dissolved organic carbon
concentration versus depth at the Alliston site.

- 49 -

ORGANIC CONTAMINANTS IN GROUNDWATERS AT SEVERAL

ONTARIO LANDFILLS

/ J.F. Barker, J. A. Cherry, D.A. Cnrey and J.P. Hewetson
Department of Earth Sciences
University of Waterloo
Waterloo, Ontario N2L 5G1

J.F. Pankow, Oregon Graduate Center, Beaverton, OR 97006

M. Reinhard, Civil Engineering, Stanford University, Stanford, CA 94305

ABSTRACT

The occurrence and migration of contaminants in groundwater impacted by
leachate at six municipal landfill sites are being studied to determine the hydro-
geological, geochemical, and microbiological influences on their mobility and persis-
tence. Of special interest are processes of attenuation (sorption, biodegradation)
and dilution (hydrodynamic dispersion). Currently, studies are underway in sandy
aquifers (CFB Borden, North Bay and Woolwich), less permeable sandy silty glacial
till (CFB Borden), and in fractured shale and dolomite (Burlington and Hamilton).
Studies at a landfill site on fractured clay-till in southwestern Ontario will begin in
1985.

At the North Bay site, where landfilling began in 1962, organic contaminants
are found throughout the 1 km sandy aquifer. Chlorinated solvents are restricted
to the immediate vicinity of the landfill, probably because of anaerobic biodegrada-
tion. Aromatic hydrocarbons may also be degraded in the final 300 m section. At
the Woolwich site, the chlorinated solvents are migrating further while aromatic
hydrocarbons are only found in low concentration even near the landfilL Aerobic
conditions appear to dominate here, resulting in persistence of chlorinated methanes
and ethanes and degradation of aromatics. Additional research into the hydrogeolo-
gical controls of cont. minant migration at these sites is underway.

At the landfill on sandy silty glacial till at Canadian Forces Base Borden,
where landfilling began in 1976, a plume of groundwater contamination has devel-
oped that is readily mappable over a distance of 150 m based on chloride and total
dissolved organic carbon. Detailed studies of the organic compounds in this plume
will soon be initiated.

Contamination of a fractured bedrock system at the Hamilton site by organics
has been difficult to define because the major contaminants are aliphatic and aro-
matic hydrocarbons which also occur naturally in these sedimentary bedrock strata.
Therefore, research into the occurrence of hydrocarbons and S-bearing heterocyclic
hydrocarbons in uncontaminated groundwaters is underway in order to identify bet-
ter "leachate indicator" parameters.

At a landfill on fractured shale in Burlington, normal inorganic indicators of
leachate contamination did not provide a basis for delineating the zone of contami-
nation. At this site 1,1,1 trichloroethane, chlorobenzene and paradichlorobenzene at
very low concentrations have been used to trace the zone of landfill impact for a
considerable distance from the landfill.

- 50 -

INTRODUCTION

It has long been recognized that municipal landfills in Ontario and elsewhere in
humid or semi-humid climatic regions produce leachate that causes contamination of
groundwater. Zones of contaminated groundwater, referred to as leachate plumes,
are commonly extensive at landfills situated on permeable, geological materials.
The hydrogeology group at the University of Waterloo began investigations of the
extent and chemical nature of leachate plumes in Ontario in 1976. Until recently
these investigations focused on the inorganic contaminants in unconfined sand aqui-
fers. The results of these inorganic studies are presented by Cherry (1982), Cherry
(1983), and Cherry et al. (1981). In the past three years our landfill investigations
have been extended to include landfills on other types of geological materials and
to include organic contaminants. The investigations of organic contaminants are
being undertaken in collaboration with the Environmental Engineering Group at
Stanford University and with the Department of Environmental Chemistry and Biol-
ogy at the Oregon Graduate Center.

Six landfills in Ontario are currently being investigated as part of this
long-term research program. Three of the landfills are situated on unconfined sand
aquifers. One of these landfills, referred to in this paper as the old Borden land-
fill, is located at Canadian Forces Base Borden near Alliston. The other two land-
fills are near the City of North Bay and in Woolwich Township near Elmira, The
landfills at these three sites are about 20 years old and the aquifers are quite
permeable so that there has been an opportunity for relatively large plumes of
leachate contamination to develop. The fourth landfill is also situated at Canadian
Forces Base Borden. This landfill, referred to as the new Borden landfill, began
operation in 1976. It is situated on a deposit of sandy silty glacial till. Because
this landfill is young and because the geological material is only moderately perme-
able, there has been much less opportunity for a zone of groundwater contamination
to develop at this site than at the other landfill sites included in this investigation.

The other two landfills are situated on fractured bedrock. One is located on
the lower part of the Niagara Escarpment in Burlington about 3.5 kilometres from
Lake Ontario. This landfill, referred to as the Bayview Park landfill, is situated on
less permeable fractured shale. The other landfill is located on the upper part of
the Niagara Escarpment in Hamilton, about 5 km from Lake Ontario. It is situated
on moderately permeable dolomite which is underlain by slightly permeable shale.
This landfill is known as the Upper Ottawa Street Landfill. Landfilling began at
these two sites between 20 and 30 years ago.

The old Borden landfill, the Bayview Park landfill and the Upper Ottawa Street
landfill are no longer in use. They have received a final cover of earth material
and the surfaces have been planted to grass. The other three landfills are continu-
ing to receive refuse.

The purpose of our current investigations at each of these landfills is to deter-
mine the nature of the inorganic and organic contamination within the contaminant
plumes and to determine the distributions of selected organic contaminants and the
processes that control these distributions. With the support of the Ontario Lottery
Fund, our investigations of most of the above-mentioned landfills will continue for
another two years. This paper is a brief report on the nature of our studies. In
addition to the landfills mentioned above, our investigations in the next two years

- 51 -

will include a landfill situated on clayey glacial till of very low permeability in
southwestern Ontario, where contaminant migration is expected to be governed by
molecular diffusion rather than by groundwater flow. A search for an appropriate
study site is near completion and field studies will begin at the selected site in the
spring of 1985.

In this paper two of the six landfills^ the new Borden landfill and the Upper
Ottawa Street landfill, are mentioned only briefly.

METHODS OF INVESTIGATION

Prior to our Investigations, a hydrogeological study of each site had been con-
ducted by consulting firms as part of the normal investigative activities that are
usually undertaken by regional governments or landfill operators. This information
was used by us in the design of our research programs, and as a basis for the
installation of a network of groundwater monitoring devices to supplement the
monitoring piezometers that were already in place at each of the sites.

At five of the study sites, monitoring networks consist primarily of multilevel
monitoring devices constructed of small diameter polypropylene sampling tubes. At
the three landfills on sand aquifers multilevel monitoring is provided by bundle pie-
zometers (Figure 1), Bundle piezometers contain eight or nine individual piezome-
ters, each extending to a different depth, to provide vertical profiles of hydraulic
head and water chemistry. No seals are used between the piezometer tips because
in cohesionless sand aquifers, the sand caves in rapidly and apparently quite tightly
around the tubes.

The bundle piezometers were installed using hollow stem augers. Their design
and use are described in detail by Cherry et al. (1983), The bundle piezometer was
developed as a monitoring device for use in investigations of landfill plumes in sand
aquifers by experimentation at the old Borden landfill and then in later years it
was applied at the Woolwich and North Bay sites.

The main advantage of the use of bundle piezometers at these sites is that
vertical profiles of water chemistry are obtainable from a single borehole. Compa-
rable profiles using conventional piezometers require many more boreholes. At
sites where the water table is deep, the efficiency of bundle piezometers is less
than at shallow water table sites because water sampling is much more time-
consuming. This was a significant problem at the Woolwich site where the depth
to the water table below ground surface ranges from 10 to 20 metres. To allevi-
ate this problem, a narrow diameter sampling pump was developed by Robin et al.
(1982) for use at this site. This pump is now manufactured and marketed by an
Ontario company (Solinst, Burlington),

The monitoring networks at the two sites on fractured bedrock consist of con-
ventional piezometers installed previously by consultants and multilevel monitoring
devices (Figure 2), The multilevel monitoring devices consist of a bundle piezome-
ter within a PVC casing. Each piezometer tube is connected to a sampling part in
the casing. The segments of borehole sampled by each tube are isolated from
above and below by an inflated packer. The packers are constructed of a chemical

- 52 -

sealant (Dowell sealant), which expands when contacted by water, which is pumped
into the PVC casing. It contacts the chemical sealant through holes in the PVC
casing adjacent to the packer. A rubber membrane covers each pacl<er to prevent
contact with groundwater surrounding the monitoring device.

The number of piezometers that can be placed in each multilevel device
depends on the size of borehole. The boreholes used at the Upper Ottawa Street
Landfill Site in Hamilton were approximately 7 centimeters in diameter. Each
multilevel device contained between 5 and 7 piezometers. At the Bayview Park
site in Burlington, the boreholes were approximately 10 centimeters in diameter so
that multilevel devices contained between ID and 12 piezometers. Because the cost
of drilling boreholes in rock is high, there is a strong financial incentive to place
as many piezometers as possible in each borehole. The disadvantage of these mul-
tilevel devices is that assembling is more difficult and time-consuming than is the
case for larger diameter conventional piezometers.

The multilevel device for bedrock was originally developed for use at the Upper
Ottawa Street landfill and later used at the Bayview Park site in Burlington.
Details regarding the design and construction of the device are provided by Cherry
and Johnson (1982), and by Cherry et al, (1985), The pump used for drawing water
samples from the multilevel devices at the fractured rock sites is the one men-
tioned above for sampling bundle piezometers in deep water table areas (Robin et
al. 1982).

At the new Borden landfill, conventional piezometers are used for groundwater
monitoring. The silty till at this site does not provide the sediment caving charac-
teristics of cohesionless sand which are necessary for the use of bundle piezome-
ters. Nor, does it provide smooth open boreholes of the type encountered in bed-
rock. Smooth open holes are necessary for the Dowell-sealant type multilevel
monitoring devices. Therefore, conventional piezometers were used at this site.

The network of conventional piezometers that existed prior to our investigations
has been augmented to provide more detailed monitoring. Nests of conventional
piezometers are used to provide vertical profiles of water chemistry. In general,
only one piezometer is installed in each borehole, the tip surrounded by a sand
pack and sealed from the rest of the borehole in the bentonite.

The monitoring network that currently exists at each of the landfill study sites
was installed in phases. A limited number of monitoring devices were installed in
each phase, from which data were acquired and used as a basis for the installation
of additional devices. The least number of monitoring sites exists at the Bayview
Park landfill site where there are 33 conventional piezometers and five multilevel
monitoring devices. The most monitoring devices exist at the old Borden landfill
where there are approximately 69 conventional piezometers and 75 multilevel
devices.

Water samples from nearly all of the monitoring devices at each of the landfill
sites have been analysed for chloride and electrical conductance. Many of the
samples have also been analysed for total dissolved organic carbon (DOC). A lesser
number of samples from each site have been analysed for a large suite of inorganic
constituents such as major ions and trace elements.

- 53 -

Trace organic compounds in selected samples from each of the sites, except for
the new Borden landfill, have been analysed by gas chromatography and mass spec-
trometry. Although many organic compounds have been identified in the plumes,
these methods provide identifications of compounds that constitute only a few per-
cent of the total mass of dissolved organic compound in any of the samples. The
sampling and analytical methods used in the investigations of organics at the Bor-
den, Woolwich and North Bay sites are described by Reinhard et al. (1984) and the
sampling and analytical methods used in the investigation of the Bayview Park
landfill site are described by Pankow et al. (1984).

In addition to water sampling, the networks of monitoring devices at each of
the landfill sites are used for water-level monitoring and for performing field tests
for hydraulic conductivity. The water level and hydraulic conductivity data enable
interpretations of the groundwater flow patterns and flow rates to be developed for
each site.

One of the objectives of these investigations is to compare the positions and
shapes of the leachate plumes determined from chemical analyses of the water
samples to the positions and shapes that one would predict on the basis of infor-
mation pertaining to hydraulic conductivity, water-table configuration and the dis-
tribution of hydraulic head. At some of the sites, mathematical models are being
used to develop more formal predictions based on these types of data. The most
recent application of a mathematical model for simulation of contaminant migration
is described by Hokkanen (1984) who successfully simulated in considerable detail
the 40 year development of the plume at the old Borden landfill.

RESULTS AND DISCUSSION

Plumes Delineated by Chloride and DOC

At the three landfill sites on sand aquifers and at the landfill on sandy, silty
glacial till, plumes of leachate contamination were easily delineated using chloride
and DOC.

The plume at the Borden landfill is the longest and widest, extending approxi-
mately 900 to 1,000 metres northward from the edge of the landfill. It is about
700 m wide with a fan-like shape. Landfilling at this site began in 1940 so that
more than four decades of contaminant migration have been necessary for the
plume to grow to this extent. The plume is thickest beneath the landfill where the
bottom of the plume goes as deep as 25 m below the water table. The plume
extends nearly to the bottom of the aquifer here because of the effect of a
ephemeral water-table mound beneath the refuse and possibly because of an effect
of plume density.

The Borden landfill plume becomes much thinner in the direction of groundwa-
ter flow, which is northward. The plume becomes thinner because the aquifer
becomes thinner in this direction. At its northern extremity, the plume exists in
only the lowest 2 or 3 metres of the aquifer. Thus, detailed vertical profiles of CI
or DOC were necessary to identify the plume in this area (MacFarlane et al. 1983).
The average northward groundwater velocity in the sand aquifer is about 10 to 20

- 54 -

m/yr in the northern-most part of the plume where the aquifer is thin. The south-
ern half or two-thirds of the plume has a relatively stable shape, whereas the
northern part is continuing to expand northward at the groundwater flow rate men-
tioned above.

The North Bay landfill is situated on unconsolidated glacio-fluvial sands gener-
ally 17 m to 25 m thick, underlain by a thin (< 2,) zone of less permeable, till and
then granitic bedrock. Contaminated groundwater flows southwest from the landfill,
under a large sand pit and discharges in springs adjacent to Chippewa Creek, about
700 m to the southwest (Figure 3).

The leachate-contaminated plume has been defined by repeatedly sampling
(1981-present) the multilevel piezometers. Figure 3 shows the contours of
maximum chloride (CI) concentration in 1982. Figure 4 shows the CI concentrations
in a vertical profile along the AA' cross section indicated in Figure 3. Whereas
the highest CI concentrations occupy the middle of the unconfined sand aquifer
near the landfill site, the high CI zone plunges to the bottom of the aquifer within
about 200 m of the landfill.

The plume that extends from the North Bay landfill does not spread laterally as
does the Borden plume (MacFarlane et at., 1983). The laterally-restricted path of
the former may be influenced by bedrock-surface control suggested by a bedrock
outcrop immediately east of the plume. This is confirmed in part, by recent geo-
physical surveys by Dr. J. P. Greenhouse of the University of Waterloo. However,
permeability variations as well as a "bedrock valley" may be limiting the lateral
spreading.

Groundwater velocities range from about 30 to about 150 m/yr with 70 m/yr
considered representative of the overall velocity. Thus, leachate has probably been
discharging near site AAA since the mid-1970's. A resampling of the piezometer
network in June, 1984 found lower maximum chloride concentrations (520 versus 840
mg/L) than in 1981/82 and the maxima occur 50-150 m from the landfill rather
than adjacent to the landfill as in 1982 (Figure 3). This displacement of maximum
CI concentrations is consistent with the 70 m/yr groundwater velocity. The
decrease in maximum CI concentration could be a result of dispersion. These
results indicate that the input of CI and, by inference, other species, from the
landfill is not constant. We will assume, however, that the ratio of organic con-
taminants to CI has been constant over time. This appears reasonable since a good
correlation of TGC (total dissolved organic carbon) with CI is obtained for most
sampling points.

The Woolwich landfill differs considerably from the other two landfills on sand
aquifers in that the refuse is deposited in pits that are well above the water table.
The bottom of the refuse and the water table are separated by 7 to 10 m of
partially-saturated sand. The thickness of the sand aquifer beneath the water table
is 15 to 20 m. The water table across the site slopes southeastward, which is the
direction in which a plume has developed since landfilling at the site began in the
mid-1960's.

The plume, delineated using Cf and DOC, extends nearly to the bottom of the
aquifer beneath the landfill. Although the plume is easily identified near the land-
fill on the basis of these two parameters, it is much less distinct beyond 300 to

- 55 -

600 m, where C\' and DOC values are much lower and are erratic in the vertical
profiles. The position of the leading edge of the plume has not been located in
detail because of these conditions.

From values of hydraulic conductivity, a southeastward hydraulic gradient and
porosity, calculated estimates of the average groundwater velocity are in the range
of about 50 to 150 m/yr. The lack of a clear indication of the plume beyond 400
or 500 m from the landfill is therefore puzzling. Investigations that are in prog-
ress are designed to shed some light on this situation.

The smallest plume exists at the new landfill at Borden where the front of the
plume has only travelled about 150 metres northward from the edge of the landfill.
Landfilling at this site began in 1976. Considering that this landfill is young and
that the hydraulic conductivity of the till is much less than the conductivity of the
sand aquifers, this plume has travelled relatively far. The average rate of travel
of the front of the plume is about 20 m, which is not much less than the average
travel rates attributed to the three plumes in the sand aquifers. The moderately
steep slope of the land surface causes lateral hydraulic gradient at the site to be
much larger than at the sand-aquifer sites.

The background concentrations of chloride and DOC at the four sites described
above are very low in comparison to the concentrations of these constituents in the
landfill leachate at these sites. Chloride and DOC are therefore well suited for
defining the extent of leachate impact on the groundwater zones at these sites.
Chloride concentrations in the leachate are generally in the range of 200 to 800
mg/L whereas the background concentrations of chloride are generally less than 10
mg/L. Background concentrations of DOC are generally less than 5 mg/L whereas
in the leachate the DOC concentrations are generally in the range of 50 to 5000
mg/L.

At the two landfills on fractured bedrock, chloride and DOC did not serve as
useful parameters for delineating the full extent of the groundwater zones impacted
by the landfill. At the Bayview Park landfill site, which is underlain by fractured
Queenston shale, the groundwater in the shale contains high concentrations of chlo-
ride derived from shale. This chloride salt is apparently a relic from the seawater
that existed when the shale and other sedimentary rocks formed in southern Ontar-
io. The natural DOC in the groundwater in the Queenston shale is generally less
than 5 mg/L. The landfill concentrations of DOC are much higher than the back-
ground values. However, the flow pattern in the shale is erratic and the landfill
concentrations appear to be quite variable and therefore DOC is not as suitable as
a leachate indicator as at the sites on the sand and on sandy silty till. DOC at
this site enables the leachate-impacted zone close to the landfill to be delineated
but it does not serve as a good indicator of the leachate plume in its down-
gradient extremity.

At the Upper Ottawa Street landfill, which is underlain by dolomite and by
dolomitic shale, the background concentrations of chloride and of DOC are high and
therefore at this site both these parameters are limited in their usefulness for
plume delineation. This site is similar to the Bayview Park site in that the
groundwater flow system is very complex due to the nature of permeability in the
fractured rocks.

- 56 -

Trace Organic Compounds In The Plumes

North Bay Landfill

Studies of trace organic occurrence and migration in the North Bay plume have
been underway since 1982, and are ongoing, in part, as a cooperative research pro-
gram with Stanford University, Results have recently been published (Reinhard et
al. 1984, 1984a) and so only a few main results are summarized here.

Figure 5 shows the vertical distribution of CI, DOC (dissolved organic carbon),
methane (CH.) and selected organic contaminants at location G adjacent to the
landfill again in 1982 (Figure 5). Whereas maximum concentrations of CI, DOC and
xylenes occur at a depth of 5 to 10 m, maximum concentrations of benzoic acid,
various phenols and trichloroethylene (TCEy) occur at a depth of about 17 m. This
distribution has been confirmed on at least four occasions since 1982.

The occurrence of TCEy, which in its industrial product form is organic liquid
denser than water, exclusively at depth suggests the possibility of a dense organic
liquid phase existing beneath this landfill which is being slowly leached by ground-
water flowing near the bottom of the aquifer. However, the much higher concen-
tration of "light" organics, benzoic acid, and phenols at the same piezometer do not
support this hypothesis. Other explanations for this vertical variation at G include
different inputs along the respective flow lines being sampled at G and possible
selective biodegradation of organic components in the different geochemical/
microbial environments along the flow lines.

(Table 1 presents the concentration of selected organics in the plume sampled
in 1982). Although many organic contaminants persist to the discharge springs, the
chlorinated methanes, ethanes and ethylenes have not been detected beyond about
200 m from the landfill (Reinhard et al. 1984 and Oct. 1984 sampling). This is
unlikely to represent retardation by sorption as more-readily sorbed organics have
moved over 700 m. It could represent only recent disposal of these organics, but
this is not considered likely given their long-term use as solvents and degreasers.
It is more likely that these organics are being microbially transformed during
migration (Reinhard et al. 1984, 1984a).

Only methanogenic bacteria have been conclusively shown to transform these
chlorinated compounds (Kobayashi and Rittman, 1982). Although methane does
emanate from the landfill in groundwater, it is not clear from CH, or CH,/C1 dis-
tributions whether methanogenesis is also occurring in the leachate plume beyond
pjezorneter G. King (1983), using the distribution of stable carbon isotope ratios
{^ C/ C) as well as the distribution of carbon between organic (DOC), inorganic
(Die) and methane (CH ) pools in migrating groundwater, indicates that methano-
genesis is occurring at least until piezometer LL (about 400 m from the landfill),
but that the final 300 m of flow might be influenced by methane oxidation. Thus
an active methanogenic environment is present where the chloroform,
1,1,1-trichloroethane and trichloroethylene are decreasing to less than 0.1 ug/1 (ppb)
levels, supporting the concept of biotransformation as the attenuation mechanism
for these organics.

- 57 -

The indication of more oxidizing conditions in the latter 300 m of the flow
system (King 1983) is interesting in view of the apparent biotransformation of some
aromatic hydrocarbons between LL and AAA (Figure 3 and Table 1). Compounds
such as 0,mandp-xylenes, 1,2,4-trimethyl benzene and napthaiene decrease signifi-
cantly, even with respect to CI, between LL-9 and AAA-5. These compounds are
generally considered degradable only under aerobic conditions. The lack of measur-
able « 0,2 mg/L) dissolved oxygen in all seriously impacted groundwaters could
indicate that the required aerobic conditions were not met, but could also indicate
that the dissolved oxygen entering these waters by dispersion (mixing) was consumed
in the transformation of these aromatic hydrocarbons.

Many other groups of organics such as aromatic acids, poly nuclear aromatic
hydrocarbons, chlorinated benzenes, are present in this plume. It is hoped that
continued research will provide information on the environments conducive to their
transformation and on the combined physio-chemical and microbial processes influ-
encing their attenuation in sandy aquifers. We view this site as an outstanding
natural laboratory in which to assess landfill-derived organic contaminant migration.
The natural complexities such as source input variation, and aquifer variability
require that conclusions be based on long-term studies with repeated sampling of
critical piezometers.

Woolwich Landfill

A total of 110 sampling points from within twenty-one bundle piezometers have
been sampled for analysis of volatile organic compounds. From this work various
organic compounds in the groundwater zone have been identified. Near the landfill,
leachate-impacted groundwater is characterized by low levels of chlorinated vola-
tiles, higher levels of aromatic hydrocarbons and phenolic compounds, and much
higher levels of carboxylic acids. These types of compounds are common in sani-
tary landfill leachate and are derived from the breakdown of organic materials and,
in some cases, from disposal of the compound itself.

The two most common and extensive of these compounds are
1,1,1-trichloroethane and trichloroethylene. There are readily identifiable concen-
trations of these compounds in the vicinity of the landfill, and much lower concen-
trations at distance southeastward of the landfill. It is possible that concentration
levels that are detectable but that are less than about one microgram per litre are
artifacts of the monitoring devices rather than actual contamination in the aquifer.
It is known that 1,1,1-trichloroethane and trichloroethylene are sometimes derived
from the plastics in glues used in such monitoring devices.

It is expected that 1,1,1 trichloroethane and trichloroethylene would exist in the
groundwater close to the landfill because these compounds are commonly seen in
contaminated groundwater at landfills in North America and Europe. Such com-
pounds are also commonly seen in groundwater where industrial pollution, not relat-
ed to landfills, occurs. The fact that these two compounds do not exist at high
concentrations beyond a distance about 300 m downgradient of the landfill is con-
sistent with the of some significant degree of adsorption in the aquifer, which
would cause them to travel less quickly than the flowing groundwater. It is also
possible that biodegradation processes cause attenuation of one or both of these
compounds in this aquifer.

- 58 -

The limit for 1,1,1 trichloroethane in drinking water suggested recently by the
U.S. Environmental Protection Agency is 200 micrograms per litre. The concentra-
tion levels in the Woolwich aquifer are very small relative to this value. The sug-
gested limit for trichloroethylene in drinking water provided by the State of New
York is 10 micrograms per litre. A few values above this limit occur very near
the Woolwich landfill but not at distances from the landfill. To our knowledge,
guidelines for these compounds in drinking water have not yet been produced by the
Ontario Ministry of the Environment. The closest drinking-water well is about 1
km from the landfill in the direction of groundwater flow. The zone of identifiable
aquifer contamination at the present time therefore has not yet moved sufficiently
far to cause closure of wells.

The persistence and mobility of organics at the Woolwich site are difficult to
define because of the limited sampling up to 1983. The results of sampling during
1984 may provide a better definition of the areal extent of organic contaminants.
Consideration of the potential long-term impact of these organics on groundwater
quality must await these results.

Borden Landfill

Groundwater from only about 20 sample points in the bundle piezometers at the
old Borden landfill were examined for organic contaminants. The general findmg
was that organic contaminants were not often present at levels much above back-
ground. Four volatile compounds were detected: chloroform, carbon tetrachloride,
trichloroethylene and tetrachlorethylene. Concentrations never exceeded 4 mg/L
(ppb) and were usually less than 1 ug/1. There was no correlation between distri-
bution of TOC or CI and these organics.

Low concentrations « 10 ug/1) and irregular distributions of aromatic hydrocar-
bons (toluene, substituted benzenes) benzothiazole, and fatty acids are consistent
with minor leaching if organic matter from the dominantly burned-fiU material.
Elemental sulphur was often found. Given a pH of about 7 and an equilibrium
among sulphur species, this would imply an Eh of about -2Qp mV _which is consis-
tent with estimates by Nicholson et al. (1983) based on SO^ -HS measurements.
Diethyl phthalate was often found, but, as this is a common piasticizer, it probably
is a contamination from sampling or piezometer material. Because of the low and
irregular organic concentrations, this site was considered to be less suitable than
the other landfill sites for additional detailed studies of organic compounds in the
plume.

Bayview Park Landfill

Organic nitrogen, phenol and dissolved organic carbon (DOC) were used as indi-
cators of the bulk organic content in groundwater. When these parameters are
present at low levels in background groundwater, the high levels contributed by a
landfill can be used to trace contaminations. DOC, is commonly used in this way
in sand and gravel aquifers where background levels are a few mg/L while those in
the plume may be lO's and lOO's of mg/L.

- 59 -

At the Bayview Park site, background concentrations of organic nitrogen, phenol
and DOC are quite low in comparison to those found in sand and gravel deposits
elsewhere in Ontario. At the three background monitoring locations, phenol has a
maximum level of 2 mg/L, while DOC values range between 2.5 and ^.0. Organic
nitrogen, the difference between total kjeldahl nitrogen and free ammonia, has a
maximum background level of 0.3.

These relatively low background concentrations provide a good contrast with
concentrations found downgradient of the landfill. Above background levels of all
tliree indicators occur in the shale for several hundred metres showing an overall
decrease with distance. Maximum levels in a piezometer beneath the landfill
establish the landfill as a major source of dissolved organic compounds.

Organic nitrogen shows the most gradual concentration reduction, with concen-
trations approaching background levels about 800 m downgradient of the landfill.
While the concentration in the refuse is almost an order of magnitude higher than
background levels, elsewhere in the shale, concentrations exceed this by only 1 or 2
mg/L.

Phenol and DOC concentrations are much more erratic than organic nitrogen
downgradient of the landfill. Within 350 m phenol levels in the most shallow pie-
zometer in each well nest are at or approaching background while levels in the
deeper piezometers are at least 50% of the maximum found in the refuse. DOC
concentrations within 600 m downgradient of the landfill are distinctly above back-
ground levels, with the exception of the deepest piezometer of this maximum dis-
tance.

Samples for trace organic analysis at the Bayview Park site were obtained using
two down-hole cartridge samplers described by Pankow et al. (198^a, 198Ab). Sam-
ple cartridges contain Tenax-GC, to which contaminants are adsorbed when the
sampler is positioned down the piezometer. The cartridge, rather than water, is
then submitted to the laboratory.

The potential influence of natural organic matter on the bulk organic concen-
tration trends necessitated the use of more diagnostic tracers of groundwater con-
tamination. Three trace organic compounds, 1,1,1 trichloroethane (TCA), chloroben-
zene (CB) and paradichlorobenzene (PDCB), were identified in groundwater and used
in this capacity. As synthetic compounds, the only source of these compounds in
groundwater is the landfill, when it is established that the piezometer materials are
not a significant source.

The distribution of TCA, CB and PDCB confirmed the vertical movement of
landfill-derived contaminants to depths at least as great as the deepest piezometer.
These compounds also indicate that there has been migration of landfill-derived
contaminants at least as far towards Lake Ontario as the farthest piezometers from
the landfill (3.5 km). Thus, the bottom and the front of the plume have not yet
been located.

All three trace organic compounds were found as far 350 m downgradient as of
the landfill. Beyond this point, CB is not detectable and PDCB concentrations are
approaching non-detectable limits. Concentrations of TCA are still at least 0.5
ug/L at a distance of 800 m, however. Although virtually nothing is known about

- 60 -

the input of the compounds to the groundwater flow system, the distance downqra-
dient to which each occurs appears consistent with their relative biodegradabiiity.

The occurrence of CB and PDCB is coincident with tritiated groundwater, which
implies input to the system after 1953. Because no landfill-related contaminants
were introduced to groundwater prior to the opening of the landfill in 1961, the
occurrence of TCA in non-tritiated groundwater suggests differential loss of these
two constituents to the shale. This loss may be due to diffusion to the shale
matrix as suggested by a free solution diffusion coefficient for tritium that is a
factor of 2.7 larger than a comparable coefficient for organic compounds structur-
ally similar to TCA. Laboratory measurements of the diffusion coefficient for
chloride and tritium in the shale are currently in progress.

SUMMARY AND CONCLUSIONS

At each of the six landfills included in this investigation, extensive zones of
leachate impacted groundwater have been identified. In each of these zones the
vertical variations of concentrations of both inorganic and organic contaminants are
large. The use of multilevel monitoring devices to determine vertical profiles of
water chemistry has been essential in the task of determining the zones where
highest contamination levels exist.

At the landfills on sand aquifers, the vertical concentration profiles of trace
organic compounds are, in general, much different than those of chloride and DOC.
Delineation using chloride and DOC therefore will not necessarily provide definition
of the most important zones of contamination with trace organic compounds. The
lateral distribution of trace organic compounds in the sand aquifers indicate that
the absorptive capacity of these aquifers is too low to prevent considerable migra-
tion of many hydrocarbons. The solid phase organic carbon content of the sand
aquifers are very low and therefore the sand provides little tenancy for absorption
of organic compounds.

Distribution of trace organic compounds at the Woolwich and North Bay landfill
sites is much more complex than that of the inorganic parameters such as chloride
and dissolved organic carbon. This suggests that much more extensive monitoring
would be required to provide a basis for predicting the future migration of trace
organic compounds relative to the monitoring detail required for inorganic contami-
nants.

By far the greatest difficulty in delineating zones of leachate migration in the
groundwater zone was encountered at the two landfill sites on fractured bedrock.
At these sites the pattern of groundwater flow is very complex because the frac-
ture network is complex. The background water chemistry also provided complexi-
ties because of high salt concentrations and more variable and abundant concentra-
tions of natural organic carbon. At the landfill site on the Queenston shale it was
necessary to use trace halogenated hydrocarbons for identifying the main extent of
the zone of leachate-impacted groundwater. At the landfill site on dolomite and
dolomitic shale a large number of inorganic and organic constituents were necessary
to provide indications of leachate contamination.

- 61 -

Our studies nt the six Inndfill sites indirnte thai wfien landfills are situRted on
rncjderately pertnunble or very permeable overburden deposits or on fractured bed-
rock, extensive zones of leachate- impacted groundwater can develop. In fractured
shale, extensive zones of contamination can develop even though the shale has a
low permeability. The potential of landfill leachate to cause groundwater contami-
nation does not seem to diminish with landfill age. Thus it is reasonable to expect
that the extent of the contaminant zones will gradually increase in future decades
and maybe even in future centuries. Although many inorganic contaminants are
present throughout leachate plumes, inorganic contaminants that are hazardous in
dririking water are rarely observed anywhere but very close to the landfills. Trace
organic contaminants pose the main threat to drinking water supplies. Fortunately,
of the six landfills that we are investigating, only one is situated in an aquifer
where water supply wells may eventually be adversely impaced by landfill leachate.

ACKNOWLEDGEMENTS

Most of the funding for the studies described in this paper has been provided
by the Lottery Fund of the Province of Ontario through the Research Advisory
Board of the Ministry of the Environment. Funds for our investigations of the
Upper Ottawa Street Landfill site were received from the Ministry of the Environ-
ment and the Ministry of Health of the Province of Ontario by way of the Upper
Ottawa Street Landfill Site Study Committee. Much of the drilling expenses asso-
ciated with the studies of the Woolwich site and the Bay view Park site was paid
for by the Regional Municipalities of Waterloo and Halton, respectively. Most of
the trace organic analyses of samples from the old Borden site, the Woolwich site
and the North Bay site were done as part of a cooperative research project
between Stanford University and the University of Waterloo funded by the U.S.
Environmental Protection Agency.

REFERENCES

Cherry, J. A., 1983. Occurrence and migration of contaminants in groundwater at
municipal landfills on sand aquifers. In; Environment and Solid Wastes, Edi-
tors, C.W. Francis, S.I. Auerbach and V.A. Jacobs, Butterworths, Boston, p.
127-147.

Cherry, J. A,, 1983. Migration of contaminants in groundwater at a landfill: A
case study. Journal of Hydrology, VoL 63, no. 1-2, May 1983.

Cherry, J. A., Barker, J.F., Buszka, P.M., Hewetson, J.P, and Mayfield, C.I., 1981.
Contaminant occurrence in an unconfined sand aquifer at a municipal landfill.
Proc. Fourth Annual Madison Conference of Applied Research and Practice of
Municipal and Industrial Waste, Sept. 28-30, Madison, Wisconsin, pp. 393-411.

Cherry, J. A., Gillham, R.W., Anderson, E.G. and Johnson, P.E., 1983. Migration of
contaminants in groundwater at a landfill: A case study. 2. Groundwater
monitoring devices, Journal of Hydrology, Vol. 63, pp. 31-49.

- 62 -

Churry, J. A. and Johnson, P.E., 1982. A multilevel device for hydraulic head mon-
itoring and groundwater sampling in fractured rock, Ground Water Monitoring
Review/, Vol. 2, no. 3, pp. ^2-A^.

Cherry, J. A., Johnson, P.E,, Blackport, R.J. and Hewetson, J.P., 1984. Development
and application of a multilevel device for groundwater monitoring in fractured
rock. Canadian Geotechnical Journal (in press).

fHokkanen, G.E.. Application of the alternating direction Galerkin technique to the
simulation of contaminant transport at the Borden landfill. M.Sc. Thesis, Uni-
versity of Waterloo, 1984.

King, K.S., 19B3. Carbon isotope geochemistry of a landfill leachate. Unpublished
M.Sc. Thesis, University of Waterloo, 120 p.

Kobayashi, H. and Rittman, B.E., 1982. Microbial removal of hazardous organic
compounds. Environ, Sci. Technol., 16, p, 171A-181A.

MacFarlane, D.S., Cherry, J. A., Gillham, R.W. and Sudicky, E.A,, 1983, Migration
of contaminants in groundwater at a landfill: a case study. 1. Groundwater
flow and plume delineation.

Nicholson, R.V., Cherry, J. A. and Reardon, E.J., 1983. Migration of contaminants
in groundwater at a landfill: a case study. 6. Hydrochemical patterns and
processes. Journal of Hydrology, Vol. 63, p. 131-176.

Pankow, J. P., Isabelle, L.M., Hewetson, J. P. and Cherry, J. A., 1984a, A tube and
cartridge method for down-hole sampling for trace organics in groundwater.
Ground Water (submitted in October, 1984).

Pankow, J. P., Isabelle, L,M., Hewetson, J.P. and Cherry, J, A., 1984b. A syringe
and cartridge method for down-hole sampling for trace organics in groundwater.
Ground Water Vol. 22, no. 3, pp. 330-339.

Reinhard, M., Goodman, N.L. and Barker, J.F., 1984, Occurrence and distribution
of organic chemicals in two landfill leachate plumes. Environ. Sci. Technol.,
(accepted for Nov, 1984).

Reinhard, M., Graydon, J.W., Goodman, M.L. and &ark®ff^<I.F., ly84a. The distribu-
tion of selected trace organics in tnr leachate^plum&Mif a municipal landfill.
Proc, 2nd Internat. Conf. Ground-WaLcr Uualify;. Res., Tulsa, \JKy March 27,
1984.

- 63 -

POLY- TUBING
- 8 mm I D
12 mm D

BINDING TAPE

EPOXY CEMENT
PLUG

j_ PERFORATED INTERVAL
WITH NYLON SCREEN

PVC PIPE
, '3mm I.D.
20 mm 0.0.

, I SLOTTED INTERVAL

WITH NYLON SCREEN

END CAP

Figure 1. Bundle Piezometer

CONVENTIONAL PIEZOMETER NEST MULTILEVEL DEVICE

BOREHOLE

SiNC- a GRAVEL
FILTER

PVC PIPE

BENTONITE SEAL

PLOTTED P1PE„
WELL SCREEN

SHALE a T LL
BACKFILL

BENTONITE SEAL

BOREHOLE

PVC PIPE CASING

POLYETHYLENE PIEZOMETER TUBES

EXPANDABLE PACKER

Figure 2. Monitoring devices installed in fractured rock.

loo 400 ti

• MULTILEVEL BUNDLE [ ■' , I , ., .1

PIEZOMETER o M 100.

Figure 3. North Bay landfill plume - areal extent of contamination based on maximum CI concentration
in (mg/1) at each bundle piezometer and location of cross section A-A'. Data from 1981.

NORTHEAST

.i^:

&■

Figure 4. Nortii bay landfill p'lume - extent of groundwater contamination along cross-section AA'
baseo on chloride con'-entration (mg/1) for 1981. Groundwater sampling points indicated
by dot 3.

- 67 -

lAHKX

poiMn
t-i s-a

DOC

>00 400

mg/l

J I I ■ 1 ■ — -J

CH.

XYLENES

too . 400 o »

J I 1 . — J

10 w

mq/t

Mfl/<

0-1 c-t

BENZOIC ACIO

Sum. phenols

TRI CHLOROCTMVLEHC

I J ■ I I I 1

O 40

noo

^g/l

looa

4000

t
^9 /I

Figure 5. florth Bay landfill plume - vertical profile of selected pafameters
at location G. Data from July 1982.

- 68 -

Table 1. Concentration of Chloride. TOC and Selected Trace Organics
in Piezometers - North Bay

0-4
(background)

4.87
25/10/81

G-5

G-9

Groundwaters
LL-9 AAA-5

Depth (m)
Sample date

5.49
20/10/83

15.07
20/10/83

13.08
20/10/83

'x- 5

20/10/83

Chloride (mg/L)

2.4

377

100

175

TOC (mg/L)

3.3

176

Volatile, Chlorinated

Organics (yq/1)

1 ,1 ,1-Trichloroethane

0.01

0.03

0.0

Trichloroethylene

0.0

1.6

0.0

Aromatic Hydrocarbons

(yg/1)

Benzene

0.3

as*

loa

3.9

Toluene

0.2

0.27

7.4

0.64

0.14

Ethyl benzene

0.1

5,4

0.03

m/p-xylene

< 0.1

0.16

o-xylene

< 0.05

4.3

2.3

0.12

1 , 2, 4-Tr1methyl benzene

< 0.05

7.8

0.21

Napthalene

< 0.05

2.7

1.5

2.7

0.0

Chlorinated Benzenes (yq/1)

Chlorobenzene

§.3

4.3

0.5

2.1

1 ,2-Dichlorobenzene

0.2

0.18

0.26

1.0

0.61

1 ,4-Dichlorobenzene

< 0.1

6.1

1 . S^

5.7

2.8

a - sampled 19/9/83

- 69 -

EPIDEMIOLOGICAL STUDY OF DISEASE INCIDENCE

AND RECREATIONAL WATER QUALITY AT SELECTED

BEACHES IN SOUTHERN ONTARIO

PROVINCIAL LOHERY PROJECT NO. 217

NANCY E. BROWN AND PATRICIA L. SEYFRIED

DEPARTMENTS OF COMMUNITY HEALTH AND MICROBIOLOGY

FACULTY OF MEDICINE, UNIVERSITY OF TORONTO

- 70 -

ABSTRACT

During the sunwner of 1983, a prospective epidemiological survey was conducted
on weekends at six Southern Ontario beaches northwest of Toronto. The study
area Included: Clalrevllle, Boyd, and Albion Hills Conservation Areas on
the Number River; Kelso Conservation Area on Sixteen Mile Creek; Heart Lake;
and Professor's Lake. A total of 9,296 persons were Interviewed at the beaches
and water samples were collected for microbiological analyses for the following
parameters: total staphylococci, fecal conforms, Escherichia coll , enterococci ,
fecal streptococci, heterotrophs, Pseudomonas aeruginosa , Campylobacter jejuni ,
Legionella sp., and viruses, preliminary statistical analysis of data revealed
that persons who entered the water (7914 of 9,296 persons I.e.: 85%) experienced
more overall Illness (p < .0001 by Fisher's exact two-tailed test), respiratory
(p < .0001), gastrointestinal (p < .0001), ear (p » .0010), eye (p » .0024),
and skin (p « .0202) problems, than those who did not enter the water. Allergenic
problems were not significantly different In the two groups (p " .2640).
Microbiological analyses of samples collected both when the beaches were open
(when epidemiological surveying was done) and when they were closed by officials
due to pollution problems. Indicated that the geometric means per 100 mL of
water were: 432 for fecal conforms and 370 for E . coll overall; 423 and
361 when the beaches were open; and 453 and 390 when the beaches were closed.
The current guideline for fecal conforms In the Province of Ontario of 100/100
mL of water was exceeded at all beaches. Investigation Is currently underway
to correlate the epidemiological and microbiological data, and to evaluate
the value of the existing guideline.

- 71 -

HmtODUCTION

Recently, much epidemiological attention associating swlnnlng-related
Illness with the bacteriological quality of recreational waters has focused
upon the work of Cabelll et a1^. (1) In the United States of America. Their
studies have primarily been undertaken on marine beaches, at various locations.
However, Jones £t a1^. (2) have Indicated that many people perceive freshwaters
to represent a greater public health risk since bacterial survival Is more
prolonged therein, and there Is a greater likelihood of Ingestion of significant
volumes of freshwater.

Whether or not bacteriological guidelines or standards are required for
recreational waters has been under debate for more than twenty-five years
(3,4). The Canadian federal government has listed the existing Canadian and
American guidelines In tabular form (5) (see Table 1). In addition, as shown
In Table 2, the Organization of the European Economic Communities (OEEC) has
Issued a directive for bathing water quality, that Is, waters wherein bathing
Is explicitly authorized by state authorities, or where bathing Is not prohibited
and Is traditionally undertaken by a 'large' number of bathers (6). No sound
epidemiological evidence exists to uphold the selection of the guidelines,
and since millions of people are utilizing bathing beaches In their free time
(7), the subject merits much closer attention, as does the selection of an
Ideal bacteriological recreational water quality Indicator or Indicators.

The prospective epidemiological Investigation conducted herein was performed
at several freshwater beaches In Southern Ontario, Canada, In order to ascertain
the role of the bacteriological quality of fresh recreational waters In swim-
ming-related illness, and to assess the value of current provincial and alternative
bacteriological guidelines for fresh recreational waters In the Province of
Ontario.

- 72 -

This report outlines the bacteriological and epidemiological analyses
performed to date. Multivariate logistic regression analysis using the SAS
(Statistical Analysis System) FUNCAT procedure (8) Is currently In progress.
This analysis Is taking potential confounders such as age, sex, race, etc.
Into account. These potential confounders were not Included in the analysis
of data by Cabelll et al^. In 1982. In addition, statistical testing of the
bacteriological data Is In progress.

- 73 -

fCTHQDS
Beach Sites

Beaches at five conservation areas, namely: Albion, Boyd, Clalrevllle,
Heart Lake, and Kelso, as well as a beach at Professor's Lake were selected
for the study (Figure 1), A detailed description of each beach Is provided
In the Appendix.
Ep1di«1o1og1ce1 Stmrey

A total of 9,296 people were Interviewed on weekends. In order to obtain
maximum numbers of beach-going subjects. Interviewers were trained uniformly
and monitored carefully. Family units were preferentially sleeted In order
to facilitate accurate and accessible follow-up Information. Beach groups
of size six or less were considered optimal. During the Initial Interview,
a contact or spokesperson was appointed for each beach group. The following
InforiMtlon was collected from each member of the beach group: relationship
to contact person, age, sex, previous swim record for the past four days,
whether the person swam or would swim on the Interview day, whether the head
was liiinersed In the water, previous or current Illness record for the past
four days, as well as the best time of day to telephone the contact person.
This Information was recorded on the "Initial Interview Form (Appendix).

For the follow-up information, an Interviewer telephoned the contact
person within seven to ten days of the Initial Interview. The interviewer
obtained answers to the questions listed on the "Telephone Follow-up Form"
(Appendix) and recorded the date when any symptoms were first noticed. The
telephone questionnaire attempted to ascertain whether Illness occurred within
three days (for reliable information, exclusion of Illnesses with long term
Incubation periods, and to attempt to avoid person-to-person spread of disease
within the household, as well as the Influence of excessive confounders) subsequent

- 74 ^

to swinning at a specified location; the synptoms of the Illness; whether
■edical attention was sought, and the physician's diagnosis; whether the disability
resulted In staying at home and the duration of the period at home; and confirm-
ation of the water exposure data. People who swam In alternate or the same
locations within four days prior to, or three days subsequent to a trial were
analyred by Instituting a separate variable.
Wcrob1o1og1c«1 Sunrey

Surface water samples were collected. In sterile bottles, at a lake depth
of 50 cm In an area with a maximum density of swimmers. At the time of sampling
air and water temperatures were noted as well as a description of the weather.
The turbidity of the water, and whether the lake was calm or wavy, were also
evaluated. Counts of the number of people on the beach who were not swimming
were done; for those in the water, assessments were made of numbers with their
head out or with their head underwater. Unless the beaches were closed for
swiming, water samples were collected twice dally, on week-ends (July 1 and
August 1 holidays Included). Closed beaches were usually sampled once dally.
A record of the sampling dates and beach status (open or closed) for the six
beaches surveyed appears In Table 3. The samples were chilled on Ice during
transport to the laboratory and processed within 6 hours.

Surface water samples were analyzed for the following parameters: total
staphylococci, heterotrophic plate count, fecal collforms, E. coll , fecal
streptococci, enterococci, P. aeruginosa , C. Jejuni , Legionella sp., and viruses.
Attempts were not made to recover Glardia sp. from the samples because weekly
screening of the telephone follow-up forms Indicated that there were no outbreaks
of giardiasis among the beach-going population surveyed.

For the enumeration of total staphylococci, surface water samples were
filtered through 0.45 urn Gelraan filters and Incubated for 24 to 48 h at 35*^C

- 75 -

on Vogel-Johnson agar (Difco) supplemented with 0.5% sodium pyruvate (R. Alico,
personal comnunication) Round, shiny black colonies were confirmed by Gram
staining in addition to Mannitol Salt agar (Difco), catalase, and coagulase
testing.

The aerobic, heterotrophic bacteria in water were enumerated on Casein-
Peptone-Starch agar plates (9) that were incubated at 20°C for seven days.

Fecal conforms and E. coli were enumerated by filtering appropriate
volumes of each sample and placing the filters on mTEC media. The plates
were placed in plastic cakettes, with two ice jars in each, and Incubated
at 44,50c (10). Fecal coliform counts were determined after 20+2 h. The
filters were then transferred to pads saturated with urea {in situ urease
test) for a 15 minute period and yellow colonies were counted (11). Yellow
colonies were verified as E. coli by oxidase and urease activities and growth
on citrate agar (Difco).

A membrane filter technique employing m-Enterococcus agar (Difco) was
carried out to enumerate fecal streptococci (12). Membrane filtration was
also used to Isolate enterococci (13). Filters were placed onto mE plates
and incubated for 48 h at 41°C. Following incubation, the membrane was transferred
to an Esculin-iron agar plate. After 20-30 min. at 41°C, small black spots
appear under positive colonies. To verify the colonies, the bile-escul in
medium of Schwan Is used in combination with: 1) growth at 45°C In BHI broth;
2) negative catalase test; 3) growth on A0% bile - blood agar; 4) positive
Gram stain; 5) acid reaction In litmus milk; and 6) esculin hydrolysis.

Pseudomonas aeruginosa organisms were enumerated by membrane filtration
using two different media for comparative purposes. The filters were placed
onto plates of a Ministry of the Environment modification of the mPA medium

- 76 -

(Appendix) and also onto mTIN medium (a new medium developed In our laboratory,
details of which are In preparation for publication). The plates were Incubated
yt 41.5^0 for a minimum of 48h.

For Isolation of Campylobacter Jejuni , lake water was filtered and the
filters were added to BNP (modified Brucella broth, see Appendix) broth for
enrichment. Flasks of broth containing the filters were Incubated under mlcro-
aerophlllc conditions (air was evacuated from an anaerobic Jar and replaced
with a mixture of N2»02» *"** CO2) at 42^0 for 4S h. The enriched culture
was streaked onto Sklrrow's agar (14), and the plates were Incubated at 42°C
for 48 h under the same atmospheric conditions as described above.

The procedure employed for the Isolation and enumeration of Legionella
penumophlla from water was developed by the Canada Centre for Inland Waters
and Is described In detail In the Appendix.
Virus Wethods

The primary concentration procedure which was used has been fully described
previously (15).

For secondary concentration, approximately IL of eluate from primary
concentration was adjusted to pH 7.0 *_ 0.2. Suitable dialysis tubing was
then filled with eluate, and both ends of the tubing were clamped securely.
The tubing was placed In plastic "cakettes", covered with polyethylene glycol
(PEG 6000) powder, and dialysis was allowed to proceed at 4^C for 18 + 2 h.
Vhen hydroextractlon was complete, the concentrate was filter sterilized through
a 0.22 urn filter, adjusted to pH 7.6, and Innoculated Into tissue culture.
The tissue culture used was: a) BS-C-1; b) RNK (Rhesus Monkey Kidney); and
c) BGM (Buffalo Green Monkey Kidney).

Cell culture lysates were processed for electron microscopy by the agar
diffusion method of direct examination (8), and by alrfuge^ ultracentrlfugatlon

- 11 ~

at 90,000 rpffi onto EM specimen grids. Samples prepared by both methods were
negatively stained with 2% phosphotungstic add (pH 7) prior to examination
In a transmission electron microscope. Samples were prepared in duplicate
by each method; at least five grid squares were examined per grid.

- 78 -

OBSERVATIONS AND RESULTS
1) Microbiology

A computer program was developed. In our laboratory. In order to calculate
bacterial geometric means. The program can be utilized to generate Information
for other such studies In the future.

Overall* geometric means of bacteriological counts were calculated based
upon the total number of samples collected, and not upon a geometric mean
of dally geometric means. For, epidemiological purposes bacterial geometric
means are calculated dally (by beach, month, and date) since the associated
epidemiological Illness rate Is ascertained dally, and thus the epidemiological
and corresponding bacteriological data are considered by date.

The overall study bacterial geometric mean data sre presented In Table
4 and Figure 2. These data reveal that for particular organisms the geometric
means for both open and closed beaches are rather similar. Closed beaches
tended to have more elevated geometric means, except for total staphylococci,
than did open beaches; however, application of statistical tests will be necessary
to ascertain whether the differences are statistically significant. At any
rate, closing the beaches appeared to have no dramatic effect In reducing
the bacterial geometric means. With respect to the current Ontario guideline
of a geometric mean of 100 fecal collforras/100 mL, and the alternative of
200/100 mL used by many others (Table 1), the guideline was surpassed for

*Eg. Over the study period, or by month, or by beach, or by status, or by
beach and status, etc.

- 79 -

both open and closed beaches separately and collectively. If one employs
a total staphylococci concentration level of a geometric mean of 100/100 mL* as
suggested by Brown (17), the level would be violated overall, and for open,
but not closed, beaches. If one utillres a geoioetric «ean concentration level
°^ ^^ E' col 1/ 100 nL, the level would have been surpassed overall, and for
both open and closed beaches. Individually. The same comments would hold
true If a concentration level of 1/100 mL was used for the potential pathogen
P. aeruginosa .

The geometric means of the bacterial counts overall, by beach, and by
beach and status (open or closed), are presented In Table 5 and Figures 3
and 4. Statistical testing will be applied later. In order to determine whether
significant differences exist between beaches, and between open and closed
beaches. At any rate. If one considers the overall geometric mean data, the
order from best to worst beach based upon fecal coll form and E. coll data
is: Professor's Lake, Heart Lake, Albion, Boyd. Clalrevllle, and Kelso.
The order, for total staphylococci becomes: Professor's Lake, Albion, Boyd,
Heart Lake, Kelso, and Clalrevllle. With respect to the geometric mean guidelines
for fecal conforms, both the 100/100 mL and 200/100 mL guidelines were exceeded
overall at each beach, at each open beach, and at each closed beach. It should
be noted that Albion and Professor's Lake were never closed. For total staphy-
lococci, the 100/100 mL geometric mean concentration level was surpassed overall
at each beach, and for open beaches (except for Albion and Professor's Lake),
and when closed at Clalrevllle, Heart Lake and Kelso, but not at Boyd. If
one utilizes a geometric mean concentration level of 100 E. coll/ 100 mL,
*Close to that reconaiended by Favero et al . (16).

- an -

the level would be exceeded overall, at open, and at closed beaches. The
same Is true If a geometric mean concentration level of 1 P. acruglnosa/ 100

mL Is utilized.

Clearly, the geometric mean guidelines and suggested concentration levels
were often surpassed at the beaches, and this Is suggestive of bacterial pollu-
tion. The value of the current guidelines Is being re-evaluated along with
the epidemiological data.
11) Epltfeiriology

In this study, a yery good response rate of 90% was attained (Table 6).

With respect to the beach population surveyed, the age distribution of
the population Is outlined In Table 7 and Figure 5, Not surprisingly, a large
proportion (88.37%) of the population was under the age of 40, and the two
largest categories, respectively, were age 20 to less than 25 (13.84%), and
age 5 to less than 10 (12.55%), with the categories of ages ten to under 15,
and 15 to under 20, following closely (both at 11.52%).

Other characteristics of the study population may be found in Table 8
and Figure 6*. A fairly good balance of males (46.15%) and females (53.85%)
participated In the study. Of those Interviewed, 28.66% were of Italian descent.
Traditionally, persons of Italian descent have represented the single predominant
type of racial background found on the study beaches herein. In addition,
most of the study population were In a high socioeconomic status group (67.73%,
persons/room ration 0.9), versus 19.71% In the middle (persons/room ratio
> 0.9 -* 1.3) group, and 12.56% in the lower group (persons/room ratio > 1.3).**

* Total numbers vary by category due to the existence of some missing values.
** Person/room ratios utillred were similar to those utilized by Cabelll et
al. (18).

- Rl -

Note also that only 35.86% of the study populatlo went Into the water (waded
OP swam) during the period four days prior to and three days after the Interview
day. With respect to food and drink consumption, both of which merit attention
In any study of this nature, 51.18% of the study population consumed home
products, 26.90% consumed both beach** and home products, 14.5% consumed
only beach products, and 7.41% did not eat or drink at the beach. In addition,
the number and percentage of Interviews, by beach, reveal that roost persons
were interviewed at Albion Hills (30.38 %) , followed by Clalrevllle (19.33%),
Heart Lake (18.43%), Professor's Lake (14.74%), and Kelso (5.16%). Closure
of the beaches was newer, at any time, within our control, and was handled
solely by public health and conservation area officials, both of whom were
acutely aware and sensitized to the political nature of the topic.

In Table 9, and Figures 7 and 8, the crude symptom rates are presented.
Persons may appear In more than one Illness category. Respiratory Illness
was defined as sore throat, cold, cough, or runny or stuffed nose. Gastro-
intestinal Illness . ^ defined as vomiting, stomachache, nausea, or diarrhea.
Skin problems were defined as boil or skin rash. Ear problems were noted.
Allergy problems were defined as allergic itch, welts, or sneering. Styes,
red. Itchy, or watery eyes comprised eye problems. Ill was defined as any
of the ailments listed on the telephone follow-up form, apart from sunburn.
Other ailments were defined as: any ailment not listed on the form, or fever,
which Is a vague, non-specific symptom, which could be put Into many categories.

** This does not Include drinking water at the beach, but that was monitored
by public health officials and there were no problem areas.

Of those Interviewed, 85.13% entered the water, that Is, waded or swam*
on the Interview day. If one compares the crude symptom rates for persons
who entered the water, versus those who did not, the following categories
of Illness were significantly elevated for persons who went Into the water:
111, respiratory, gastrointestinal, and other categories (all of which had
a Fisher's exact two-tailed test p-value of < .0001; ear (p- .0010); eye
(p» .0024); and skin (p= ,0202) problems. Allergy was not significantly different
between the two groups (p= .2640).

A finer classification of exposure revealed that 14.87%, 7.15%, and 77,98%
of the study population respectively, 1) did not enter the water, 2) were
waders and 3) were swirnuers. Using the chl-square test, there were significant
Illness differences between the three groups for 111, respiratory, and gastro-
intestinal (p= .0001) categories, other ailments (p« ,0008), ear (p« ,0058),
skin (p= .0060), and eye (p= .0108) problems, but not for allergic (p« ,4217)
problems. A dose-response relationship (that Is, no water < wader < swlnroer)
can be seen for the 111, respiratory, gastrointestinal, eye, ear, allergy,
and other categories, but this was not upheld In the skin category.

At present, multivariate logistic regression analysts (upon data from
which missing values have been removed, In order to permit comparison) is
In progress. The analysis is designed to determine the relationship between

Uaders were those people who only entered the water to knee depth at most.
Swlnmers were those people who Immersed their bodies further than waders.

- 83 -

the bacteriological and Illness data*. In the presence of potential confounders.
The statistical package being utilized to analyze the data Is the SAS FUNCAT
procedure (8). The SAS LOGIST procedure (19) was found to be less useful
In handling discrete parameters, whereas the FUNCAT results were found to
be comparable to those generated by the GLIM system (20).

- 84 -

1. Cabelll. V.J., Dufour, A. P., McCabe, L.J.» and Levin, N.A. 1982. Swlnnlng-
assoclated gastroenteritis and water quality. An. J. EpIdeMlol. 115:
606-616-

2. Jones, P., and White, W.R. 1984. Health and amenity aspects of surface
waters. Wat. Pollut, Control M: 215-225.

3. Stevenson, A.H. 1953. Studies of bathing water quality and health.
Am. J. Public Health f3: 529-538.

4. Moore, B. 1959. Sewage contamination of coastal bathing waters In England
andWales: a bacteriological and epidemiological study. J. Hyg. 57: 435-472.

5. Health and Welfare Canada, 1983. Guidelines for Canadian recreational
water quality. Health and Welfare Canada, Ottawa, Canada. H46-20/1983E.

6. Commission of the European Communities. February 1976. Council Directive
of 8 December 1975 concerning the quality of bathing water. Official
Journal of the European Conmunltles pp. 31/1 - 31/7. 76/160/EEC*

7. Shuval, H.I. 1975. The case for microbial standards for bathing beaches.
In: Discharge of
London, pp. 95-101.

In: Discharge of Sewage from Sea Outfalls . Ed. H. Gameson. Perganon,

8. SAS. 1982. The FUNCAT procedure. SAS User's guide; Statistics . SAS
Institute Inc., Cary, North Carolina, pp. Z57-Z85.

9. Staples, D.G., and Fry, J.C. 1973. A medium for counting aquatic hetero-
trophic bacteria In polluted and unpolluted waters. J. Appl . Bact. 36:
179-181. ~

10. Pagel , J.E., and Vlassoff, L.T. 1979. Determination of performance charac-
teristics for fecal collform enumeration procedures. Abstracts of the
Annual Meeting of the American Society for Microbiology: 229.

11. Dufour, A. P., Strickland, E.R., and Cabelll, V.J. 1981. Membrane filter
method for enumerating Escherichia coll . Appl. Environ, Microbiol* 41:
1152-1185. ~

12. American Public Health Association. 1971. Standard Methods for the
Examination of Water and Wastewater, 13th edition. American Public Health
Association, Inc., New York.

13. Levin, M.A., Fischer, J.R., and Cabelll, V.J. 1975. Membrane filter
technique for enumeration of enterococci In marine waters. Appl.
Microbiol. 30: 66-71.

14. Sklrrow, M.B. 1977. Campylobacter enteritis: a "new" disease. Br. Med-
J. 2: 9-11.

15. Health and Welfare Canada. 1981. A Study of disease Incidence and recrea-
tional water quality In the Great Lakes. Phase 1. 81-EHD-67.

- 85 -

16. Favero, M.S«, Drake, C.H.. and Randall, G.B. 1964. Use of staphylococci
as Indicators of swlnilng pool pollution. U.S. Public Health Service
Public Health Reports 79: 61-70.

17. Brown. N.E. 1983. The bacteriology and epidemiology of swlmlng-related
Illness. M.Sc. thesis. University of Toronto.

18. Cabelll. V.J., Dufour, A. P., Levin, M.A., McCabe, L.J., and Habernan,
P.W. 1979. Relationship of Microbial Indicators to health effects at
marine bathing beaches. A.J.P.H. 69: 690-696.

19. SAS. 1983. The Loglst procedure. SUGI Supplementary Library User's
Guide . 1983 Edition. SAS Institute Inc. pp. 181-202.

20. Baker, R.J. , and Nelder, J.A. 1978. The GHm System (Release 3) . General 1 red
Linear Iterative Modelling . Royal Statistical society, Herts, England.

Table 1. Summary of Maximum Limits for Collfoms In Primary Contact Recreational Water Presented by Various Agencies

Agency

1. (U.S.) National
Technical Advisory
Coml ttee

2. Province of British
Columbia

3. Inland Waters
Directorate*
Department of the
Envlronoient

4. Alberta Environment

5. OnUrio Ministries of
Health, Environment

6. Connlttee of the
Great Lakes Upper
Mississippi River
Board of State
Sanitary Engineers

7. World Health
Organization

8. U.S. Environmental
Protection Agency

9. Saskatchewan
Environment

10.

Sampling Regime

Not less than 5 samples
taken over not more than
a 30-day period

Total Collfonns

Not less than 5 samples
taken over not more than
a 30-day period

At least 10/30 days

Manitoba Clean

Environment
Commission

Not less than 5 samples
taken over not more than
a 30-day period

Not less than S samples
taken over not more than
a 30-day period

Not less than 5 samples
per month

{500/100 bL

Geometric mean
{1000/100 bL

Geometric mean
{1000/100 hL

Geometric mean
{1000/100 nL

Fecal Collfoms

{200/100 mL, nor shall
more than lOX of the
samples exceed 400/100 mL

Running geometric mean
{200/100 ml, nor shall
more than lOX of the
samples exceed
400/100 mL

{200/100 mL

Geometric mean
<2W/100 mL

Geometric mean
< 100/ 100 mL

Geometric mean
{200/100 mL. No
sample to exceed
1000/100 mL

Ev coH {1000/100 nL

Geometric mean
{200/lX mL nor shall
nwre than lOX of the
samples exceed
400/100 nL

Geometric mean
{200/100 mL

Median <200 HPN/100 mL

Other

Fecal

streptococci,
PseudoBonas .
Staphylococcus

Fecal

streptococcus.
Pseudomonas

Median <500 MPN/
100 mL

(from Health and Welfare Canada. 1983; 5)

Reference

National Technical
Advisory Comnittee, 1968

Province of British
Columbia, 1969

Environment Canada, 1972

Alberta Environment, 1977

Ontario Ministry of Health
1975; Ontario Ministry of
the Environment. 1978

Committee of the Great
Lakes Upper Mississippi
River Board of State
Sanitary Engineers, 1975

World Health Organization.
1975

U.S. Envlrormental
Protection Agency, 1976

Saskatchewan Environment,
1977

Province of
Manitoba, igyg

Table 2. Quality Requirenw«tts for Bathing Water

Parameters

Minimum

sampling

frequency

r^thod of analysis
and Inspection

Microbiological:

Total conforms /lOO ml

500

10.000

Fortnightly
(1)

Faecal conforms /100 ml

100

2,000

Fortnightly
(1)

Fermentation In multiple
tubes. Subculturing of the
positive tubes on a confirm
at1on medium. Count accord'
ing to MPN (most probable
number) or membrane filtra-
tion and culture on an
appropriate medium such as
Tergltol lactose agar, endo
agar, 0.4% Teepol broth,
subculturing and 1dent1f1ca
t1on of the suspect coloniei

In the case of 1 and 2, the
Incubation temperature Is
variable according to
whether total or faecal
conforms are being
investigated.

Faecal
streptococci

/lOO ml

100

(2)

Litsky method. Count
according to MPN (most
probable number) or
filtration on membrane.
Culture on an appropriate
medium.

Salmonella

/I litre

(2)

Concentration by membrane
filtration. Inoculation
on a standard medium.
Enrichment - subculturing
on isolating agar-
identification.

Entero

viruses PFU/10 litres

(2)

Concentrating by filtration
flocculation or centrifugl*
and confirmation.

Physico-chemical:
pH

6 to 9 (0)

(2)

Electrometry with
calibration at pH 7 and 9.

Colour

No abnormal
change in
colour (0)

Fortnightly
(1)

(2)

Visual inspection or
photometry with standards
on the Pt. Co. scale.

*G = conforms if 901 of samples (801 for total coliforms and
fecal coliforms) correspond to specifications.

♦*I ^ confonns if Q'i'K nf <;amnlp«; corresnond to soecifications-

(from Commission of the European Communities. 1976; 6)

I able L - cuntiriueu

- 88 -

Parameters

•G

Minimum

sampling

frequency

Method of analysis
and inspection

Mineral oils mg/litre

^0.3

No film
visible on
the surface
of the water
and no odour

Fortnightly
(1)

(2)

Visual and olfactory
Inspection or extraction
using an adequate volume
and weighing the dry res

Surface-active mg/litre
su bs tances ( T auryl -
reacting with sulfate)
methylene blue

^ 0.3

No lasting
foam

Fortnightly
(1)

(2)

Visual inspection or
absorption spectrophoto-
metry with methylene blu

Phenols mg/litre
(phenol CrHrOH
Indices) * *

^0.005

No specific
odour

s< 0.05

Fortnightly
(1)

(2)

Verification of the absc
of specific odour due tc
phenol or absorption
spectrophotometry 4-
amlnoantlpyrine (4 AAP)
method.

Transparency m

1 (0)

Fortnightly
(1)

Secchl's disc.

Dissolved oxygen
X saturation O2

120

(2)

Winkler's method or
electrometric method
(oxygen meter).

Tarry residues and
floating materials such
as wood, plastic
articles, bottles,
containers of glass,
plastic, rubber or
any other substance.
Waste or splinters

Absence

Fortnightly
(1)

Visual Inspection.

Ammonia mg/litre NH

(3) "

Absorption spectrophoto
metry, Nessler's method
indophenol blue method.

Nitrogen Kjeldahl

mg/litre N

(3)

Kjeldahl method.

Other substances
regarded as Indications
of pollution

Pesticides mg/litre
(parathion, HCH,
dieldrin)

(2)

Extraction with appropriA
solvents and chromatogre^
determination

iflDie c, VfOntinuea

- 89 -

Parameters

17 Heavy metals such as:

- arsenic mg/Utre As

- cadmium Cd

- chromium VI Cr VI

- lead Pb

- mercury Hg

Cyanides

mg/lltre

Nitrates and mg/lltre NO3

phosphates

po:

Minimum

sampling

frequency

(2)

Method of analysis
and Inspection

Atomic absorption possible
preceeded by extraction

Absorption spectrophoton^
using a specific reagent

Absorption spectrophotomelrf
using a specific reagent

G ■ guide.
I ■ mandatory.

(0) Provision exists for exceeding the limits in the event of
exceptional geographical or meteorological conditions.

(1) When a sampling taken In previous years produced results
which are appreciably better than those In this Annex and
when no new factor likely to lower the quality of the
water has appeared, the competent authorities may reduce
the sampling frequency by a factor of 2.

(2) Concentration to be checked by the competent authorities
when an Inspection in the bathing area shows that the
substance may be present or that the quality of the water
has deteriorated.

(3) These parameters must be checked by the competent authorities
when there Is a tendency towards the eutrophlcatlon of the
waters.

Table ;; continued.

- 90 -

Article 5

1. For the purposes of Article 4, bath-
water shall be deemed to conform to the
relevant parameters:

if samples of that water, taken at the
same sampling point and at the Intervals
specified In the Annex, show that 1t
conforms to the parametric values for
the quality of the water concerned, in
the case of:

- 95X of the samples for parameters
corresponding to those specified In
column I of the Annex;

- 90t of the samples In all other cases
with the exception of the "total
coHfonn" and "faecal coliform"
parameters where the percentage may
be 80%

and if. In the case of the 5, 10 or
20% of the samples which do not comply:

- the water does not deviate from the
parametric values In question by more
than 50%, except for microbiological
parameters, pH and dissolved oxygen;

- consecutive water samples taken at
statistically suitable intervals do
not deviate from the relevant
parametric values.

2. Deviations from the values referred
to in Article 3 shall not be taken into
consideration in the calculation of the
percentage referred to in paragraph 1
when they are the result of floods,
other natural disasters or abnormal
weather conditions.

Article 6

1. The competent authorities in the
Member States shall carry out sampling
operations, the minimum frequency of
which is laid down in the Annex.

2. Samples should be taken at places
where the dally average density of
bathers is highest. . Samples should
perferably be taken 30 cm below the
surface of the water except for mineral
oil samples which shall be taken at
surface level. Sampling should begin
two weeks before the start of the
bathing season.

3. Local Investigation of the conditions
prevailing upstream In the case of fresh
running water, and of the ambient
conditions In the case of fresh still
water and sea water should be carried

out scrupulously and repeated periodically
in order to obtain geographical and
topographical data and to determine the
volume and nature of all polluting and
potentially polluting discharges and
their effects according to the distance
from the bathing area.

4. Should Inspection by a competent
authority or sampling operations reveal
that there Is a discharge or a probable
discharge of substances likely to lower
the quality of the bathing water,
additional sampling must take place.
Such additional sampling must also take
place if there are any other grounds
for suspecting that there Is a decrease
in water quality.

5. Reference methods of analysis for the
parameters concerned are set out In the
Annex. Laboratories which employ other
methods must ensure that the results
obtained are equivalent or comparable

to those specified in the Annex.

Article 7

1. Implementation of the measures taken
pursuant to this Directive may under no
circumstances lead either directly or
indirectly to deterioration of the
current quality of bathing water.

2. Member States may at any time fix more
stringent values for bathing water than
those laid down in this Directive.

Article 8
This Directive may be waived:

(a) In the case of certain parameters
marked (0) In the Annex, because

of exceptional weather or geographical
conditions;

(b) when bathing water undergoes natural
enrichment in certain substances
causing a deviation from the values
prescribed in the Annex.

Table 2 continued. - 91 -

natural enrichment means the process whereby,
without human Intervention, a given body
of water receives from the soil certain
substances contained therein.

In no case may the exceptions provided for
in this Article disregard the requirements
essential for public health protection.

(from Commission of the European Conmunltles, 1976; 6)

Table 3. Record of the Sampling and Closure Dates for the Six Beaches Surveyed In 1983.

BOYD 1

CLAIREVILLE

ALBION

HEART

LAKE

lSO

PROFESSOR'S LAKE

Date Sampled

Open

Closed

Open

Closed

. Open

. Closed

, Open

Closed

Open

Closed

Open

Closed

June 25

June 26

July 1

July 2

July 3

July 9

July 10

July 16

July 17

July 23

l£3

July 24

July 30

July 31

Aug. 1
Aug. 6
Aug. 7

X
X

X
X
X

X
X

Aug. 13

Aug. 14

Aug. 20
Aug. 21
Aug. 27
Aug. 28

X
X
X

X
X
X
X

X
X
X

X
X
X
X

X
X

- 93 -

Table 4. Overall Geometric Means of Bacterial Counts

Geometric Neans

Open and Closed
Beaches

Open
Beaches

Closed
Beaches

Organi sin

Pseudomonas aeruginosa (mPA)

172

118

Pseudomonas aenicilnosa (mTIN)

172

118

Heterotrophs

2799063

171

2589361

118

3328985

Enterococci

172

118

Total staphylococci

12S

163

142

109

Fecal streptococci

100

172

118

131

Escherichia coll

370

172

361

118

390

Fecal conforms

432

172

423

118

453

lauic 3. ucwneinc neans or Bacterial Counts by Beach
i) For open and Closed Beaches

Beach

^7 aeruginosa

*«

= Boyd C.A.*

= Clairevllle C.A.

- Albion Hills C.A.

= Heart Lake

= Kelso C.A.

= Professor's Lake

39
31
44
34
8
16

(mPA)

P. aeruginosa

(mTIN)

18
8
4
6

30
4

11
6
4
5

17
3

Heterotroph!

6033695
2407375
3174590
3216177
1540471
412746

EnterococcT

37
72
19
39
77
22

Total
staphylococci

103
282

93
176
180

Fecal
streptococci

CA - Conservation Area.

* For toUl stapftylococci n « 37, 28, 42, 32, 8, 16

11) For open Beaches

^. aerugl

152
203

55
120
125

E. coH

443
794
298
224
1147
163

Beach

Inosa

10
18
44
24
6
16

P. aeruginosa

— (srnn —

8
9
4
7
27
3

Heterotrophs

16323929
4563167
3174590
2279226
1100521
412746

Enterococci

Total
staphylococci

29
46
19
27
116
22

391
502

93
217
160

Fecal
streptococci

}

. coll

** For total stapf^ylococcl n - 8, 15, 42, 22, 6, 16

272
200

55
116
244

1102
624
298
251

1619
163

Feca
collfo

503
967
327
268
1283
216

Fecal
conforms

1275
763
327
300

1799
216

111} For Closed Beaches

Beach

P. aeruginosa

Heterotrophs

Enterococci

Total
staphylococci

Fecal
streptococci

E. coll

Feca

n****

(mPA)

(mTIN)

collfo

4280896

125

324

365

993127

131

145

206

1107

1343

8056505

110

127

170

204

4224926

255

408

465

****: For heterotrophs n - 29, 13, NA, 9. 2. NA.

so
I

- 96 -

Table 6. Response Rates Achieved In the Epidemiological Study

Interviewed subjects

No. persons Interviewed 10287

No. completed Interviews 9296

No. incompleted Interviews 991

% success 90.37

*No answer, wrong number, not-In-service.

- 97 -

Table 7. Age Distribution of Beach Population
Categories No.

-> under 5

5 * under 10
10 - under 15
15 - under 20
20 - under 25
25 - under 30
30 - under 35
35 - under 40
40 - under 45
45 - under 50
50 - under 55
55 - under 60
60 - under 65
65 - under 70
70 - under 75
75 - under 80
80 - under 85
85 - 90 and over

842

9.06

1167

12.55

1071

11.52

1071

11.52

1287

13.84

988

10.63

997

10.73

792

8.52

423

4.55

248 <

2.67

172

1.85

0.99

0.87

0.37

0.18

0.10

0.03

0.02

9296 100

- 98 -

Table 8. Other Characteristics of the Study Population

I) Sex (n ■ 9270) Ho^

Male
Female

II) Race (n - 9250)
lUllan
Other

III) Socioeconomic status* (n ■ 9296)
High (persons/room ratio <0.9)
Middle (persons/room ratio >0.9 4=1.3)
Low (persons/ room ratio >1.3

4278

46.15

4992

53.85

9270

100

2651

28.66

6599

71.34

9250

100

No.

6296

67.73

1832

19.71

1168

12.56

9296 100

!r-'

♦Ascertained from contact person, and taken to be representative of associates
1n the same beach group, because of sensitivity of the question and duration
of ascertalrment. This Index Is similar to that used by Cabelll et al- 1979.
Note that the kitchen(s) and bathroom(s) are not counted as rooms.

1v) In water during period 4 days

before or 5 days after Interview day (n' 92B7) No.

Yes
No

v) Food and Drink Consumption (n « 9228)
No food or drink
Home products
Beach products
Both home and beach products

No.

3330

35.86

5957

64.14

9287

100

No.

684

7.41

4723

51.18

1339

14.51

2482

26.90

9228 100

- 99 -

Table 6. continued

v1) Interviews completed at each beach (n ■ 9296)
Beach 1 » Boyd Conservation Area
Beach 2 - Clalrevllle Conservation Area
Beach 3 - Albion Hills Conservation Area
Beach 4 » Heart Lake Conservation Area
Beach 5 - Kelso Conservation Area
Beach 6 « Professor's Lake

No.

1112

11.96

1797

19.33

2824

30.38

1713

18.43

480

5.16

1370

14.74

9296 100

- 100 -

^iiK»

Table 9. Crude Symptom Rates

1) By water (wade or sw1m). no water

Water (n- 7914) No water (n- 1382) P- value*

(X)

111

7.53

2.03

<.0001*

Respiratory
Gastrointestinal

Sk1n

3.08
2.10
0.88

0,72
0.43
0.29

<.0001*

.0202*

Eye
Ear

0.97
0.95

0.22
0.14

.0024*
.0010*

Allergy
Other

0.47
1.53

0.22
0.29

.2640
.0001*

(Sunburn - not considered
an Illness) (10,64)

(5.64)

(<.0001*)

ii) By wade, sw1m, no water
Symptom (n= 9296) Wader (n" 665)

111

Respiratory

Gastrointestinal

Skin

Eye

Ear

Allergy

Other

3.91

1.35

1.40

0.15

0.60

0.45

1.20

(Sunburn - not Illness) (8.87)

Slmner (n= 7249)

7.86
3.24
2.20
0.95
1.01
0.98
0.47
1.56

(10.80)

No water (n'^ 1382) p^yalue

Itl —

2.03

.0001*

0.72

.0001*

0.40

.0001*

0.29

.0060^

0.22

.0108^

0.14

.0058^

0.22

.4217

0.29

.0008^

(5.64)

(.0001

a = By Fisher's ^xact two-tailed test,
b = By Chi -square.

I
o

>-<
r

HJLLSL -

10000m ^' '"

HEART UUcin

gLAIREVILLEl

^MILTON%

Location of Conservation and
Recreation Areas

Figure 1. Location of Sample Sites

* Metropolitan Toronto Regional Conservation Authority,

- 102 -

1000J^^^'"®tr'>c means
I wv/cof counts for

& closed

ibeaches

Legend

• Open ( nsl18)
oCloeed (n-54)

♦ Overall (n -172)

100 r

Figure 2. Overall Geometric Means of Bacterial Counts

- 103 -

lOOOuGewnetric means
luwutpf counts for

beach

Legend

• Boyd (n> 39)
oClairevilie(n-31)

• Albion (n«44)

• Heart Lake (n.3b)
aKbIso (niB)

" Professor b
Lake (n>16)

5 ""

£T 9 ff O

S Z

ii 3

CO
CD

Figure 3. Geometric Means of Bacterial Counts by Beach

- 104 -

1000

100-

Legend

So*

i I'
I II

3.*;

Figure 4. Geometric Means of Bacterial Counts by Beach and Status

- 105 -

Age Distribution of
Beach Population

s
2

(0
11.

Figure 5. Age Distribution of Beach Population

+• , 2 "

SEX SOCIOEC FOOD

ONOMIC CONSUMED

RACE STATUS OTHER

SWIM*
EXPOSURE

BEACH

100
90
80
70
60
50
40
30

0»

% of
population

1 1

3 S

3x1

^ A
O O

Figure 6. Other Characteristics of the Study Population
* or water i.e. re: wade or swim

a 3

Symptom rates for people who did
and did-not enter the water

W— 7914
NW-1382

tlL

ru PL ri-

WNW WNW

S WNW WNW WNW WNW WNW

(0
UJ

Figure 7. Crude Symptom Rates for People who Entered the Hater (Waders and Swimmers) and
Those Who Did Not

5
4

3
2
1

Symptom Rates
for Waders ,

Swimmers and

People who did
not enter the
water

n ^

ILL

8
7
6
5
4
3
2
1

U^rfL

$MO ^wo $(oa

AU.ERGY OTHER SUNBURN

W n.665
S n*7249
D n »1382

-n^ H-L r-n

^(OQ ^(OQ ^(OO ^WQ ^(OQ
RESPIRATORY \ SKIN EYE EAR

GASTROINTESTINAL

Figure 8. Crude Sympton Rates for Waders » Swlnriers. and Those Who Did not Enter the Water,

109 -
APPENDIX

ClalrevlUe Conservation Area (Humber River to Lake Ontario)

This dawned up portion of the Humber River offers a wide expanse of
open swimming area. A cut grass, wooded, park, serviced by dirt roads
surrounds the swimming area. Facilities Include: boating, picnic
tables, parking, washrooms, a snack bar. and fishing areas In streams
found within the area. There Is, however, a charge for entering this
area. A large population of Canada geese and local birds were often
observed. The narrow, consistent, sandy beach consists of medium grain,
sandy particles. The roped-off swimming area Is shallow, but eventually
drops off. and the sediment Is sandy with occasional small stones. The
area Is often frequented by large organized groups. At this site
washrooms are located at a great distance from the beach.

Bo yd Conservation Area (Hurtiber River to Lake Ontario)

Nearby to ClalrevlUe Conservation area, and just south of the town
of Klelnburg Is Boyd Conservation Area, for which a charge Is levied
to enter the parkland. A huge cut grass, wooded, park, with numerous
trails and streams, serviced by excellent finished roads. Facilities
^ Include: picnic tables, parking, washrooms, a snack bar. barbeques. and

nature trails. As with ClalrevlUe Conservation Area, large groups
often frequent the area, especially those of Italian background. A
narrow, medium grain, consistent, sandy and stoney beach, and roped-off
swinmlng area, are well sheltered by the surrounding tall trees and
large hills. The swinming area is fairly shallow and the sediment is
quite stoney.

- 110 -

Albion Hills Conservation Area (Humber River)

This conservation area lies on the upper part of the Humber River,
north of the Town of BoUon. A beautiful cut grass, large «f00ded and
hilly park, serviced by paved roads surrounds the SMlmnlng area.
Facilities Include: picnic tables, parking, washrooms, a snack bar,
and a fishing area. A charge Is levied for entering the area. The
narrow, consistent, sandy, u-shaped beach sonslsts of medium grain,
sandy particles. The roped-off swimming area Is shallow for a long
y way off shore, but eventually drops off, and the sediment Is sandy.

Heart Lake Conservation Area (Heart Lake)
(on Heart Lake Road North)

This lake is the result of a danned up spring, and is located near

Brampton. A large rolling park in which one descends using steps from

the parking areas to swimming areas, is found therein. Facilities

include: picnic tables, parking, washrooms, a snack bar. a. fishing

area, and a boating area. The narrow, consistent sandy beach consists

of medium grain sandy particles. The surrounding area is roped off

and has a wery soft, mushing sediment. The swimming area drops off,

)

in depth, fairly quickly. This area, unfortunately, suffers from

algal blooms frequently.

Kelso Conservation Area

The Kelso Conservation Area was acquired to construct a dam and
reservoir on the west branch of the Sixteen Mile Creek. The dam Is an
earth fill structure with a concrete spillway and stilling basin.

- Ill -

Two gates on top of the dam regulate the water levels and flows. The
reservoir. Glen Eden Lake* has a storage capacity of 1,190 acres • feet,
a surface area of 85 acres, and a maximum depth of 35 feet. As a
recreational area. Kelso Is very popular, open year round, and visitors
pay to enter the area. Sumner faclltles. Include a beach area, change house,
snack bars, washrooms, group camping areas, boating facilities, and
picnic areas. Over 200 acres of the Niagara Escarpment are within the
area. A large, medium grain, consistent, sandy beach Is a great attraction.
^ The swimming area drops off gradually.

Professor's Lake

Professor's Lake near Brampton (further south than Heart Lake; on
Torbram Road North) Is a very beautiful lake produced by damning up a
gravel quarry. A charge Is levied to enter the area. Facilities Include:
picnic tables, parking, washrooms, a snack bar. a water slide, and a
boating area. The area Is kept Imnaculately clean. The narrow, consistent
sandy beach consists of medium grain sandy particles. The swlirmlng area
is roped off and has a sandy sediment. The swimming area drops off
) gradually.

^ Bum-up

Eu3 CooMrvit
araa

TON-ElS

CLAIREVILL

CONSERVAT
AREA

Figure 9. A Map of Clalrevllle Conservation Area

I
to

BOYD

CONSERVATION
AREA

Figure 10. A Map of Boyd Conservation Area

-4^55'

ALBION HILLS
CONSERVATION

AREA

BES Cofw«r¥ttlon Ar«i
QZa Built-up land
==: Road

Figure U. A Map of Albion Hills Conservation Area

79*47-5'

- 43 45

HEART LAKE
CONSERVAOPN

AREA

/ \

/ ^

SJ'^ss

Bustf I enlarged!

TOILETS

'^4/f^

^ , ^ BEACH

-AiA2-s!

BRAMPTON

SSSnznnmm

Figure 12. A Map of Heart Lake Conservation Area.

Kelso joa built-up

Conservatlonlcss conservati
Area

Figure 13. A Nap of Kelso Conservation Area.

^'mlco--^6reelc\a

Figure U, A Map of Professor's Lake

TX) QL Wn ISM

•CM aou

ZSOTflSCM

SDOI

85Km

9MN TSU

(SOK) 1DBIO

MDMOi

SBnOUINI

MWWTiWI IBdA '<r JUO

IVKBe xg ,1.)

■osMU ivnsai

«T

>X]K> a39BdCMUCD XX&

MCZTaX -XA

NXM TXDi MO *wam *A

com

)l#(eH JO ngu

(c^ 'T<40 ai ff

Ill I-.-, m-n ~ rm-i

MQi MXABUa 'KOXia

- SIT -

- 119 -

i±

rrm ---i 1 1 1 1 1 1 1 1 1

IBBVt

I -

-MD

MDTMnJOUKC

mhlmk.

Tl 11

Ml I

III 1

Mil

Tl 1 1

HVb

Z. Smk en cottACt dv?

In vatar lo or
en contact dqr

Betf tfdv?

A«t wdid («Mt la)
oi ocBtaet da/

3dVB7

s-eat.

At thla baach
only? (Yaa or Hd)

(Ma or No)
1 3

XZI.SMaUaf aiy watar

Did not aat food or
dtrlnk at baach

Bat food taous^ at
baadi (or driidc)?

Eat fiood, at baarh,
brou^ £icB hcaa
(or dElnk)

flHSICM

d svmo

6 - zaaa

Ed dOy i

!(r 3 day

a afbarl

Lta inwqin.fint notiaad)l.a. J\l^

A. Soea throat

B. Mwr

C. Paid or wa^

D. Ikmy or atuffiad

E. Earadia« Itory Eara

F. Styaa or zad« itciv
or watary ayaa

G. Stoadi adia or

a. DLarxhaa<

I. Vadtlna

J. BoUa

K. SIdn raah
L. AUaz^nU LtAt
walta or laatlng

K SwibuznT

lA raa

«lth aiadlar
(not at bMch)

P> Odd you aae doctor?

2-2, 3-3

Q. IB<yioaAa? Giva in 2ad bowT

Did you ati^ at hcne
of Ulnasa?

tf. Ha* many day?

^. I£ Ulnaaa aarioua, may wa hava

your doctor'a nan* «id phcna mtfarT

- 120 -

P. aeruginosa Isolation Medium niPA (Ministry of the Environment Modification)

L - lysine monohydrochlorlde 5.0 g
Sodium chloride

Yeast extract 2.0 g

Xylose 2.5 g

Sodium thiosulfate 5.0 g

Sucrose 1.25 g

lactose 1.25 g

Magnesium sulphate 1.5 g

Ferric ammonium citrate 0.8 g

Sodium desoxycholate 0.1 g

Phenol red 0.06 g

Agar* 15.0 g

Distilled water (sterile) 800 ml

♦Add ingredients, except agar, to sterile beaker. Adjust pH to 7,60 and
then add agar. Heat to 93°C until agar starts to dissolve. Do not
autoclave. Cool to 70°C, check surface pH (5 ml of mPA in a small plate)
Adjust pH If necessary. Cool to 50-55®C (no higher than 60®C). Add
antibiotics, mix thoroughly, and pour Into petrl dishes. Final surface
pH 7.1 + 0.1.

Antibiotics

Sulfapyrldlne 0.088 g

Kanamycin sulfate 0.0085 g

Nalidixic acid 0.037 g

Actldlone 0,15 g

Dissolve in 200 ml sterile H,0. Heat gently to above 50^C.

- 121 -

BNP Broth

Brucella broth (Gifoco) was supplemented with 0.2S Fe So^.7H20 and
autoclaved at 121^C. The broth was cooled to 55^C and supplemented
with vancomycin 15.0 mg/L; Trimethoprin 7.5 ng/L; Polymyxin B 30*000
lu/L; sodium bisulfate 0.025S; and sodium pyruvate 0.05t.

- 122 -

DUKRATICM and ISOUTIOli OF lEGIOWEUA WgUMOPHIlA FROM IMTEK BY

DIKCT PUTING PROCEDURE
(CanMto Cffitrv for Inland Haters)

Sample Collection

Filter 1000 ml water sample or 100 ml sewage sample through one or more Miftrane
filters. With a sterile 10 ml pipette 10 ml of sterile filtrate Is rowved
and placed Into an ointment Jar.

The membrane filter Is placed Into the ointment Jar, bacteria containing side
down.

Sample Processing

Place capped ointment Jar In sonic sink for 10 min. to dislodge bacteria from
membrane filter.

1 ml of filtrate-bacteria mixture Is removed and placed In capped test tube.
Add 1 ml pH 2.0 buffer, mix and let stand for 10 min.

Add 1 ml neutrallzer, mix and using 1 ml pipette, dispense 0.1 ml, 0.3 ml
and 0.6 ml of neutralized mixture onto the surface of 3 predrled labelled
BCYE agar plates.

Using glass spreader, spread Inoculum over the surface of the BCYE agar.
After the surface of the agar has dried. Invert the petrl dishes, place In
plastic bags, seal and place In V^C Incubator for 4 to 5 days.

Screening Plates

Count all grey-vfhite flat colonies (3-4 mm) with opaque centres, often having
ground glass appearance.

Select all or typical representative colonies and with wire loop transfer
each colony to

a) BCYE agar 370C, b) XYE agar 20^0

c) Tryptone Soya Agar 37°C d) Tryptone Soya Agar 20'*C

Incubate the 37^ plates for 3 days and the ZQ<^C plates for 7 days. (Each
plate may be divided Into 4-6 sections to accorvnodate more colonies.

Only those colonies growing on BCYE agar Incubated at 37°C should be considered
as potential Legionella pneumophila .

Calculations

The total nurter of Legionella pneumophia per original sample nay be
derived from the following formula:

- 123 -

T ■ N X ( ^ ) X 10. Wh«re T ■ total niMber bacteria per org. sanple volume
^ N ■ number of typical colonies per plate

A ■ total volume of pH treated sample
B " total volume of pH treated sample plated

- 125 -

VOLATILIZATION RATES OF ORGANIC CHEMICALS
OF PUBLIC HEALTH CONCERN

T.P. Halappa Gowda and J.O. Lock
Water Resources Division
Gore & Storrie Limited
1670 Bayview Avenue
Toronto, Ontario
M4G 3C2

- l^t) -

VOLATILIZATION RATES OF ORGANIC CHEMICALS
OF PUBLIC HEALTH CONCERN

T.P. Halappa Gowda and John D. Lock
Water Resources Division, Gore & Storrie Limited
Toronto, Ontario

ABSTRACT

A study of volatilization of organic chemical compounds of public health concern
from streams and rivers has been carried out. The volatilization rates are
expressed in terms of liquid film coefficients (K[_). The' K|_ values of
ethylene gas are determined from data on ethylene and rhodamine WT dye
concentration distributions which were collected in provincial streams and
rivers of differing hydraulic characteristics.

Relationships among Kl, channel hydraulic parameters and chemical properties
of organic compounds have been developed through statistical analyses as well as
dimensional analysis using the Buckingham ir-Theorem. The relationships for
ethylene gas, derived from the dimensional analysis procedure, have been found
to provide better correlations with the observed K|_ values. A sensitivity
analysis of various parameters on Kl has also been carried out. A
statistical relationship developed by Rathbun and Tai (1982) has been found to
underestimate Ki_ values for the provincial streams and rivers. An
evaluation of the theoretical relationships between K|_ and molecular
properties (critical volumes and molecular weights) using experimental data
reported in the literature has revealed that the relation between <[_ and
critical volume is suitable to calculate Kl values for highly volatile
organic chemcials using the known Kl values of ethylene gas or other
compounds (eg. propane, benzene); however, for moderate and low volatile
compounds, no relationships could be obtained fom the available data.

The computation of Kl for a given organic compound using the relationships
developed in this study is outlined in a step-by-step procedure. An example
illustrates the various computations involved. In general, the method is
applicable to volatile compounds with Henry's law constants greater than
10"^ atm-mVmol .

- 127 -

INTRODUCTION

Assessment of potential public health and other risks of organic chemicals in
streams and rivers requires detailed evaluations of their transport, cycling and
fate within the aquatic environment. Therefore, considerable importance is being
given to gathering data on various physical, chemical and biological processes and
factors related to these organic chemicals of public health concern. One of the
important processes common to many of these organic compounds is volatilization.
It is particularly necessary to know the volatilization rates of chemical compounds
when dealing with emergency situations such as accidental spills of such compounds
in order to predict the transport, fate and public health risks of the spills.

The volatilization process represents the physical transport of organic compounds
through the water-air interface into the atmosphere. And, hence, there is a
similarity between the widely-known atmospheric reaeration and the volatilization
processes. Recently, Mackay and Yuen (12) and Rathbun and Tai (17) have shown that
the volatilization rates applicable to organic compounds can be estimated from the
field data gathered for the measurement of reaeration coefficients using dye and
ethylene as tracers.

Volatilization rate, also known as "desorption rate", is a coefficient which is
analogous to the first-order rate coefficient, expressed in the units
day"^ However, the liquid film or mass transfer coefficient (expressed
in m/day), is generally utilized in the volatilization studies since it is a more
fundamental quantity (12, 17).

Volatilization of organic compounds has been studied through laboratory experiments
(2, 4, 7, 15, 16) as well as by making use of data on tracers from natural streams
and rivers (12, 17). A modified tracer technique originally developed by Rathbun,
et al (18) for measuring stream reaeration coefficients Involves the use of a
desorbing gas (eg. ethylene or propane) and a conservative material (eg. rhodamine
dye) as the tracers. Mackay and Yuen (12) and Rathbun and Tai (16, 17) have shown
that the data gathered by the modified tracer technique are also useful for
volatilization studies of organic compounds; and Rathbun and Tai (17) have
obtained statistical relationships between channel hydraulic parameters and liquid
film coefficients. Field studies using the modified tracer technique (with
ethylene gas and rhodamine WT dye as the tracers) have also been carried out by the

- 128 -

Water Resources Branch, Ontario Ministry of the Environment, Toronto, during
1978-80 in streams and rivers of differing hydraulic characteristics. These data
have been utilized to estimate the liquid film coefficients and to evaluate their
relationships to channel hydraulic parameters and molecular properties of organic
compounds. The detailed results of the study are presented elsewhere (3). The
objectives of this paper are as follows:

1. To develop relationships among liquid film coefficients, bulk-flow
characteristics of river channels and chemical properties of ethylene gas.

2. To evaluate the relation of liquid film coefficients to critical volumes and
molecular weights of organic compounds, and to determine their suitability to
compute liquid film coefficients of other organic compounds.

THEORETICAL ASPECTS
Review of Basic Concepts

The volatilization of an organic compound from the water phase to the air phase is
dependent on the physical and chemical properties of the compound, the presence of
other chemical compounds, the hydrodynamic and other physical properties of the
water body, and the physical properties of the atmosphere above the water surface
(7, 9 - 11). The physical and chemical properties of organic compounds affecting
volatilization include molecular diameter, molecular weight, Henry's law constant
and diffusion coefficient. The liquid film coefficients are also influenced by the
presence of some modifying materials including adsorbents, electrolytes, emulsions
and organic films (2, 11). The hydrodynamic and other physical properties of the
river channel include flow rate, width, depth, velocity, bed slope, bed roughness,
turbulence level and wind-induced currents. Factors such as suspended sediment
concentration and the presence of other chemicals in the water phase could also
affect volatilization. The atmospheric properties of concern include wind speed,
stability and other factors. Temperature affects the vapour pressure and
solubility, and thus, influences volatility of chemicals.

- 129 -

Theoretical concepts of volatnization of chemical ccinpounds from water to the
atmosphere have been presented by various researchers (for eg., see Ref. 7-10, 23,
24). A review of these theoretical approaches and their limitations can be found
elsewhere (9). The two-film theory of Lewis and Whitman (7) is generally utilized
to describe the volatilization of organic compounds from water. The two-film model
assumes that the bulk air and water phases are uniformly mixed, and that these two
phases are separated by thin films of air and water, as shown schematically in
Figure 1. The main resistance to gas transport is considered to exist in the
liquid and gas phase interfacial layers (or films). Within these two films, the
transport of the exchanging gas is assumed to take place by molecular diffusion.
Then, application of the Fick's first law of diffusion for gas transport through
each layer results in

F = kL (CsL - Cl) = kG (Cq - Csq) W

in which F = the flux of gas; kL and kg are the mass transfer coefficients
for the liquid and gas phases, respectively (meters/day); Cl and Cq are
the concentrations in the bulk liquid and gas phases, respectively; and Csl
and CsG are the liquid-phase and gas-phase concentrations at the interface,
respectively (see Figure 1). The mass transfer coefficients (also termed "exchange
constants") are defined by the general relation

k = Dfn/6 (2)

where Dm is the coefficient of molecular diffusion of gas in the appropriate
layer of thicknesses (see Figure 1). The coefficients, kL and kQ, are
measures of the flux of gas per unit concentration gradient.

Following Cohen, et al (2), the transport across the two-layer system shown in

Figure 1, can also be expressed as follows on the assumption that the

concentrations immediately on either side of the interface are in equilibrium as
expressed by a Henry's law constant, H;

F = Kl (Cl - Cs) (3a)

where Cs = Py/H (3b)

and J = J -^ RL (^^5

Kl kL HkQ

- 130 -

FIGURE 1 - SCHEMATIC REPRESENTATION OF TWO-FILM
GAS TRANSFER MODEL

- 131 -

in which Kl is the overall mass transfer (or liquid film) coefficient based on
the liquid phase (meters/day); R is the gas constant (atm-mVmol -K); T is
the temperature (°K); Py is the atmospheric partial pressure (std, atm.);
and Cs is the concentration in the liquid phase in equilibrium with Py (or
saturation concentration).

The resistances to mass transfer are given by the reciprocals of the mass transfer
coefficients. Therefore, Eq. 3c can be written as:

rx = ri_ + rQ W

where rj is the overall resistance; and ri_ and rg are the resistances

offered by the liquid and gas films, respectively. Then, the fractions of the

resistances to mass transfer in the liquid and gas films are given by

n 1 + (i^g/'^l)

re = 1 (5b)

rj I -^ (i^L/i^G)

where ri_ = _! (^^)

and re = RT/HkQ (5d)

Equations 5a-b can be utilized to determine the relative magnitudes of the

resistances to mass transfer in the liquid and gas phases, respectively. The

resistance may be dominant in either the liquid or the gas phase, or in

phases, depending on the relative magnitudes of ki_, kg and H.

both

From Equation 5a. the smallest % of resistance in liquid film is seen to occur for
the largest kL and the smallest kQ. Rathbun and Tai (17) utilized the
reported maximum value of kt = 5.77 m/day from the data on benzene,
chloroform, methylene chloride and toluene, and kQ = 480 m/day based on
evaporation of water in a canal ( corrected for the different molecules using
the square root of inverse ratio of molecular weights), and obtained resistances
in liquid film for ethylene and propane to be 99. 69?^ and 99. 9U, respectively, at
25°C. Based on this analysis, the resistance to volatilization of ethylene and
propane from water is mostly in the liquid film. Rathbun and Tai (17) have also
evaluated the percentage of resistance in liquid film as a function of the Henry's
law constant by utilizing the data on k|_, kg and H for organic

- 132 -

compounds reported in the literature. Their results show that, for average
conditions, more than 90% of the resistance is in the liquid film for compounds
with H values of about 10"^ atm*mVg mole.

Relation of K| to Molecular Properties

The molecular properties of chemical compounds which affect Ki_ values Include
molecular weight, molecular diameter and diffuslvities (9, 15, 17, 23, 24). The
relationship of molecular weight. M, to Kl 1s expressed as

K|_ <x M"°'^ (6)

The relationship of molecular diameter, d^, to Kl Is given by

Kl « dn,"^ P)

Rathbun and Tai (17) cite several studies wherein the molecular diameters have been
used successfully of adjusting the Kl for a tracer gas to an organic compound.
Often, molecular diameter for an organic compound of interest is not readily
available; and hence, the diameter Is estimated from critical volume or the longest
dimension of a model of the compound. The coefficient, Kl, has also been
related to the molecular diffusivity of organic compounds in water by

in which 0^ = the molecular diffusion coefficient at a given temperature of
river water; and n is an exponent. Rathbun and Tai (16) utilized data on
volatilization for a number of chlorinated hydrocarbons presented by Oilling (4) in
conjunction with 0^ values calculated from an equation presented by Hayduk and
Laudle (5) to obtain a value of n = 1.19; however, the 9S% confidence limits were
±0.64. Roberts and Dandliker (21) found n = 0.66 for six chlorinated and
fluorinated hydrocarbons. Other works cited by Mackay and Yuen (13) report values
of n to be 0.5 and 0.67.

Equations 6-8 can provide the basis for calculating the values of Kl for given
organic compounds from the known values for tracer gases. Since M and d are basic
properties of organic compounds. Equations 6 and 7 can be applied for adjusting the
Kl values. However, the molecular diffusivity. Dm. is dependent on the
properties of river water as well as chemical compound; and hence Equation 8 may

- 133 -

not adequately describe the effect of Dm on <!_. Thus, it is reasonable to
conclude that Equations 6 and 7 are applicable for adjusting the Kl value of
one compound to another, whereas the effects of Dm on Kl need to be
considered in a broader context in the analyses.

Relation Between Kl Values of Tracer Gases
And Organic Compounds

The ratio, R, between Kl values of two compounds is written as (15, 19, 23, 24)

KlG (9)

R =

Klorg

in which Klq and Klqrg ^^^ the liquid film coefficients of a tracer gas and
organic compound, respectively. Rathbun, et al (19) and Rainwater and Holley (15)
have presented values of R between various tracer gases based on laboratory
experimental studies. Also, the values of R between three tracer gases (oxygen,
propane and ethylene) and a group of four organic compounds have been obtained by
Rathbun and Tai (16) through correlations of Kl values.

Recently, Rainwater and Holley (15) have presented theoretical aspects related to
the use of molecular diameter and critical volume in determining the values of R.
The relationship can be written in the form

0.9H2 .0.311.

KlG / dORG V-''' /VcORG\ '

Rv = = ( ) = (10)

KlORG \ dG / \ Vcg /

in which Ry is the ratio based on critical volumes; dQ and dgRG
denote, in order, the molecular diameters of tracer gas and organic compound; and
VCQ and VcORG ^"^^ ^^^ critical volumes of the tracer gas and oganic
compound, respectively. This relationship Is based on the assumption that the
molecules are spherical, and hence, the critical volume is proportional to the cube
of the diameter. Since the exponent 0.942 appearing in Equation 10 is very close

- 134 -

to unity. Rainwater and Hoi ley (15) suggest that the relationship between R and the
ratio of molecular diameters is in reasonable agreement with Equation 7. If
molecular weight is considered to be the basis for adjusting <(_ values, then
from Equation 6, the following expression is obtained:

,, / u X . 5

/ "org \
\ mg /

Rm = = 1 I (11)

klorg

in which Rm is the ratio based on the molecular weights, and Mq and
MqrG denote, in order, the molecular weights of tracer gas and organic
compound. Detailed evaluations of Equation 10 and 11 for adjustment of K|_
values of tracer gases to organic compounds and their limitations are presented
later.

The Rathbun-Tai Relationships for Natural Streams

Rathbun and Tai (17) have correlated Kl values at 20''C for ethylene and
propane with hydraulic data from seven streams and rivers. The total numbers of
Kl values utilized in the analyses were 54 and 53 for ethylene and propane,
respectively. Different forms of predictive equations were evaluated using a
normalized root-mean-square (RMS) error as the criterion of comparison. The
experimental Kl values ranged from 1.68 to 7.54 m/day for ethylene and 1.13 to
6.94 m/day for propane. The range of channel hydraulic parameters were -
discharge: 0.047 - 5.95 mVs; depth: 0.101 - 0.555 m; velocity: 0.050 -
0.439 m/s; and bed slope: 0.0538 - 0.631 m/km. The correlation relationships
obtained by Rathbun and Tai (17) are as follows:

Ethylene: Kl = 146 Z°'^^° (US)''"^^'^ (12)

Propane: Kl = 141 Z^"^^^ (US)""^^^ (13)

in which Z = mean depth (m); U = mean velocity (m/s); and S = bed slope (m/m).
The RMS errors associated with Equations 12 and 13 were 25.6% and 34.8%,
respectively. Since these relationships are empirical in nature, they are subject
to limitations in their applicability to other streams and rivers.

- 135 -

Relationships from Laboratory Studies

Mackay and Yuen (13) measured volatilization rates of 11 organic compounds of
varying Henry's law constants (5.57 x 10"^ to 5.18 x 10"^ atm.
mVmol at 20°C) in a wind-wave tank in the laboratory. The experiments
were carried out at six different windspeeds in the range 5.96 - 13.2 m/s. From
these studies, the authors have obtained equations for K|_ as a function of wind
velocity and Schmidt number (defined by the ratio, kinematic
viscosity/diffusivity). Since their relationships do not include such factors as
depth and velocity of water, they have a limited scope for application to natural
streams and rivers.

Southworth (25) has developed a method for estimating mass transfer coefficients of
polycyclic aromatic hydrocarbons having Henry's law constants in the range
10"^ - 10"* atm-mVmol. The method makes use of laboratory
data. The relationships developed include depth, velocity, windspeed and molecular
weight; however, they do not include other factors such as channel bed slope,
viscosity and molecular diffusivity. Therefore, the relationships are also subject
to some limitations in estimating liquid film coefficients.

Dimensional Analysis

The application of dimensional analysis as a tool of developing relationships
between dependent and independent variable parameters in hydraulic studies is well
known (22). It involves a systematic organization of the variable parameters into
the smallest number of significant dimensionless parametric groups. Such an
organization is facilitated by the fact that a mathematical relationship must be
dimensional ly homogeneous. The dependence of K|_ on various parameters can be
written in the form

Kl = f(B,Z,U,LS,g,p,u.Dn,) (14)

in which B = channel width; L = length of reach; g = acceleration due to gravity;
p = density of water at T°C; and n = absolute viscosity of water.

- 136 -

An omission from Equation 14 is the wind velocity, which is known to affect <(_;
however, because of a lack of data on windspeeds for the field studies considered
herein, the parameter has not been included in the dimensional analysis.

In Equation 14, there are nine parameters which are to be varied independently.
According to the Buckingham 7T-Theorem (22), these can be grouped into six
dimensionless ir-terms. Numerous combinations of the prameters into the
dimensionless grouping is possible. For this study, the following functional
relationship has been selected:

m- 'Hirifn^-fn^rifr (^^)

in which ao is a numerical constant (dimensionless); aj - as are the
exponents of the individual dimensionless groups; and v = kinematic viscosity
defined by the ratio of absolute viscosity to density (i.e., u/p). Each
dimensionless term appearing on the right hand side Is related to a characteristic
property affecting K|_, as follows:

8 = Aspect ratio of channel
I

LS = Ratio of drop in height to channel depth
~Z (related to energy loss)

uZ = Rn = Reynolds number of flow indicative of viscous effects

u^ =: F = Froude number denoting gravitational effects

\^ = (^\ (^ \" '^nSc = product of Reynolds (R^) and
Dfn \m ) (^"Omj Schmidt (Sq) numbers

DATA COLLECTION AND ANALYSES

The modified tracer technique developed by Rathbun et al (18) was utilized by the
Ontario Ministry of the Environment to gather data in shallow streams and rivers.
The details of this technique have been presented elsewhere (1). The field
procedure involved simultaneous injection of ethylene gas and rhodamine WT dye

- 137 -

solution (as a line source across the channel) for a known duration of time at the
head of a study stretch of a river channel, and monitoring for dye and ethylene gas
concentrations as a function of time at selected downstream stations as the plume
passed those stations. These data are utilized to calculate the volatilization (or
desorption) coefficient, Ky, for a channel reach using the relationship:

((Cg/Cd)uS ) (16)

Kv = 1 sin -^ ^ '

t ((Cg/Cd)dS )

in which Cg and cj = concentration of ethylene gas and dye, respectively;
us and ds = subscripts denoting upstream and downstream stations, respectively; t
= travel time in the channel reach (days); and Ky = desorption coefficient of
ethylene gas (day"\ base e) at the average river water temperature.

The liquid film coefficient for ethylene gas in each reach is then determined from
(16)

Kl = Kv Z (17)

Field data were collected from five streams and rivers of varying hydraulic
characteristics. Figure 2 shows schematic layouts of the various study segments
of the rivers and streams.

The field survey dates for various streams and the salient hydraulic
characteristics of the stream channels are presented in Table 1. The channel bed
slopes were obtained from topographic maps. Generally, the stream channels are
very shallow, the channel bed being composed mostly of rocks, pebbles and sand.
Three surveys were carried out in the Grand and Speed rivers and one survey was
conducted in each of the other streams.

An inspection of the channel hydraulic data presented in Table 1, shows that the
channel widths, depths and velocities are identical for the following survey
conditions:

Grand River, June 11, 1978 - Reaches AB and BC
Speed River, June 7, 1978 - Reaches BC and CO
Speed River, August 24, 1978 - Reaches BC and CD

- l^H -

[IPLHJMCNIAL CHANNEL

INJECTION POINT

STN 13

AVON RIVER

^.,.— — '

^ INJECTION POINT

■

{

BADEN CREEK

^ . . 30Q

METRES

•J-'"^

V--

, INJECTION POINT

"^^^

NETM RIVrR

300

METRES

INJECTION POINT

S£FFn RIVFR

FIGURE 2 - SCHEMATIC DIAGRAMS OF STREAM AND RIVER SEGMENTS

- 139 -

TABLE 1
SUMMARY OF CHANNEL HYDRAULIC DATA

STREAM

SURVEY

DATE

REACH

L
(m)

(mVs)

(m)

Z
(m)

U
(tn/d)

(m/m)

AVON R.

9/04/80

148.00

0.110

7.31

0.108

12009.6

0.0034

127.25

0.110

7.23

0.092

14256.0

0.0034

BADEN CR.

7/27/77

274.00

0.045

3.11

0.118

10627.2

0.0063

1180.00

0.045

3.47

0.236

4752.0

0.0063

GRAND R.

6/02/78

2555.00

4.090

29.47

0.680

17280.0

0.0013

GRAND R.

7/11/78

115.00

4.270

29.69

0.685

18144.0

0.0013

2240.00

4.270

29.69

0.685

18144.0

0.0013

2555.00

4.270

29.69

0.685

18144.0

0.0013

GRAND R.

8/23/78

2440.00

7.290

34.18

0.790

23328.0

0.0013

NITH R.

8/03/77

303.00

0.153

17.91

0.305

2419.2

0.0006

260.00

0.153

13.80

0.264

3628.8

0.0006

SPEED R.

6/07/78

1480.00

3.360

30.68

1.090

8726.4

0.0011

1720.00

3.360

38.60

0.440

17280.0

0.0011

650.00

3.360

38.60

0.440

17280.0

0.0011

2380.00

3.360

38.60

0.440

17280.0

0.0011

SPEED R.

8/10/78

1480.00

2.120

29.84

0.890

6998.4

0.0011

1720.00

2.120

38.60

0.403

12009.6

0.0011

SPEED R.

8/24/78

1480.00

2.430

30.09

0.943

7430.4

0.0011

1720.00

2.430

38.60

0.414

13392.0

0.0011

650.00

2.430

38.60

0.414

13392.0

0.0011

2380.00

2.430

38.60

0.414

13392.0

0.0011

- 140 -

Therefore, in these three cases, the two successive reaches have been combined into
one reach (i.e.. Grand River - Reach AC; and Speed River - Reach BD), The
hydraulic characteristics for these cases are given in Table 1 for the individual,
as well as the combined reaches.

The liquid film coefficients, <[_, at the average instream temperatures,
calculated from Equations 13 and 14, are tabulated in Table 2. The values are seen
to range from 0.97 to 8.07 meters/day. The river water temperatures in various
streams ranged from 16° to 25°C. It should be noted that the data for the
Grand River - June 11, 1978, and Speed River - June 7, and August 24, 1978, surveys
Include the Ki_ values for the individual reaches as well as the combined
reaches.

The development of a relationship between Ki_ and the channel hydraulic
parameters, by the dimensional analysis (Equation 15), requires data on densities
and viscosities of stream waters as well as molecular diffusion coefficient of
ethylene gas. Dm, for the instream temperature conditions. The densities,
p, and absolute viscosities, y, were obtained from the CRC Handbook of
Chemistry and Physics (27). The diffusion coefficients were computed from (5, 9)

13.26 X 10"' (18)

^r. =

T' 1 - 1 »» w ■ 5 8 9

where 0^ is the molecular diffusion coefficient of ethylene gas (cmVsec);

u is the absolute viscosity of water (centipoise or 10"^ g/cm sec);

and Vb = 44.4 cmVmol Is the molar volume of ethylene gas computed by the

LeBas method (9). The values of density, absolute viscosity and molecular

diffusion coefficient are given In Table 2.

RESULTS

Relations from Dimensional Analysis

The values of various dimensionless terms of Equation 15 were calculated by using
the data summarized in Tables 1 and 2. The IMSL routine RLSEP was utilized to
determine the regression coefficients Aq - as appearing in Equation 15.
From this procedure, the following relationship is obtained:

- 141 -

TABLE 2
VISCOSITY, DENSITY, DIFFUSIVITY AND LIQUID FILM COEFFICIENTS

STREAM

SURVEY REACH T

DATE

{°C) (Kg/m-day) (Kg/m"*) (m^/day) QT^C @20OC

(m/day) (tn/day)

AVON R.

9/04/80

20.7

90,74

998.0559

0.000116

2.66

2.61

20.7

90.74

998.0559

0.000116

2.39

2.35

BADEN CR.

7/27/77

25.0

76.90

997.0448

0.000140

3.57

3.17

25.0

76.90

997.0448

0.000140

3.04

2.70

GRAND R.

6/02/78

19.0

88.73

998.4052

0.000119

1.80

1.84

GRAND R.

7/11/78

17.0

93.40

998.7748

0.000112

8.30

8.92

17.0

93.40

998.7748

0.000112

4.29

4.60

17.0

93.40

998.7748

0.000112

4.64

4.98

GRAND R.

8/23/78

22.0

82.50

997.7704

0.000129

8.07

7.69

NITH R.

8/03/77

25.0

76.93

997.0448

0.000140

2.63

2.33

25.0

76.93

997.0448

0.000140

2.26

2.01

SPEED R.

6/07/78

16.0

95.82

998.9430

0.000109

0.97

1.07

16.0

.95.82

998.9430

0.000109

3.81

4.19

16.0

95.82

998.9430

0.000109

1.95

2.14

16.0

95.82

998.9430

0.000109

3.01

3.31

SPEED R.

8/10/78

16.7

94.12

998.8252

0.000111

3.10

3.35

16.7

94.12

998.8252

0.000111

3.74

4.05

SPEED R.

8/24/78

19.0

88.73

998.4052

0.000119

5.73

5.87

19.0

88.73

998.4052

0.000119

3.31

3.39

19.0

88.73

998.4052

0.000119

5.27

5.39

19.0

88.73

998.4052

0.000119

4.33

4.44

- 142 -

1 •♦ • 1 6

-1-89

K,.3.Z(B) (.^) (UZ)-(yi)"-»(UZ,--' (^,,

V ' ^ gZ / \ D

The various statistical parameters associated with this correlation relationship
are presented in Table 3. These statistical characteristics of the regression fit
indicate that Equation 19 satisfactorily describes the functional relationship
between K|_ and the channel bulk flow parameters and chemical characteristics
for the range of data presented in Tables 1 and 2. The F-statistic values
associated with individual dimensionless terms of Equation 19 were in the range
0.56 - 4.73, the values for the last three terms being significant In the range
6 - 13%; however, the significance levels for the first two terms exceeded 20%.
Therefore, a regression fit was carried out by considering the last three terms of
Equation 19 to obtain the following relationship:

»Ci. = 7.16(yZ )--(U^)---(UZ)---' (20)

U \ V / \gT ' ^Dm '

The statistical characteristics of the regression fit are given in Table 3. which
indicate a satisfactory correlation relationship. The F-values associated with the
individual terms were in the range 2.44 - 6.18 which correspond to significance
levels of 3-5% approximately. Thus, Equation 20 also appears to describe the
functional relationship satisfactorily. This relationship can be utilized when
data on channel slopes are not readily available.

Equations 19 and 20 can be simplified to obtain the following relationships:

Kl = 2.70 g °-'^ V ^-"^ Dm ''^^ B°*'* 1-°-^^ u"-'^ (LS)^-^^ (21)

Kl= 7.16 g °-^' V -'-'' Dm'-'' Z*"-*' U""**' (22)

- 143 -

Table 3
CORRELATION CHARACTERISTICS OF RELATIONSHIPS

Statistical Parameter

Equation 19

Equation 20

Variation explained by the fit

66.6t

62.14t

RMS error

6.2*

5.94*

Overall F-value

3.59

6.02

Significance level

4.6«

1.1*

- 144 -

According to Equation 21. Kl would decrease with an increase in depth, whereas
Equation 22 indicates Kl increasing with depth. Based on the known
relationships of reaeration coefficient to channel hydraulic parameters, the
exchange rates at the air-water interface tend to decrease with increasing depths
(due to lower turbulence levels) and vice versa. Thus, 1t is thought that Equation
21 is superior to Equation 22.

Equations 19 to 22 are dimensional ly homogeneous, and hence, the units of Kl
are dependent on the units of various parameters appearing on the right hand side.
For example, if the various parameters are in the gram-meter-second units, then
Kl is in meters/second. Note that these relationships are specifically valid
to estimate Kl for ethylene gas at a given temperature.

In order to evaluate the predictive capabilities of the relationships, the values
of Kl for ethylene gas for the streams and rivers listed in Table 1, were
computed from Equations 21 and 22. The results are presented in Table 4 along with
the observed Kl values, the latter being in the range 0.97 - 8.07 m/day. The
values obtained from Equation 21 are in the range 1.87 - 5.57 m/day. whereas
Equation 22 gives values in the range of 2.51 - 6.31 m/day. Figures 3 and 4 show
plots of the computed versus the observed Kl values. The average Kl values
are 3.46 m/day for the observations, and 3.03 and 3.28 m/day from Equations 21 and
22. respectively. The ratios (predicted KL/observed Kl) vary from 0.44 to
2.11 with Equation 21. and 0.51 to 2.64 with Equation 22;and the average ratios are
1.01 and 1.12, respectively. These results indicate that the predictions of
Equation 18 and 19 compare favourably with the observations. On an overall basis,
the predictive ability of Equation 21 is slightly better than Equation 22.

The applicability of the relationships obtained from dimensional analysis to other
streams is evaluated by utilizing the data on liquid film coefficients at 20°C,
Klo, presented by Rathbun and Tai (16). The values of Klq for four oganic
compounds (viz. benzene, toluene, chloroform and methylene chloride) were found to
be approximately the same, based on laboratory studies. These Klq values are
related to the Kl values for ethylene by Klq = 0.753 Kl- The river
hydraulic data given by Rathbun and Tai (16) include water depth, velocity and
discharge; but the bed slopes and reach lengths are not reported. Thus, these
data are suitable to compute the liquid film coefficients for ethylene from
Equation 22. Table 5 shows the hydraul ic data for the streams as well as the

- 145 -

TABLE 4
COMPARISON OF OBSERVED AND COMPUTED K|_ FOR ONTARIO STREAMS

OBSERVED

(m/day)

EQUATION 21

(m/day) K

EQUATION 22

(m/day) K

EQUATION 12

K K

LC LC

(m/day) K

2.66

2.10

0.79

2.51

0,95

2.01

0.75

2.39

2.22

0.93

2.66

1.11

2.06

0.86

3.57

3.34

0.94

4.17

1.17

2.61

0.73

3.04

3.01

0.99

3.19

1.05

2.26

0.74

1.80

3.80

2.11

4.22

2.35

2.85

1.58

4.64

3.14

0.68

3.65

0.79

2.92

0.63

8.07

5.57

0.69

6.31

0.78

3.44

0.43

2.63

2.57

0.98

2.43

0.92

0.64

0.24

2.26

2.60

1.15

2.86

1.27

0.74

0.33

0.97

1.87

1.93

2.57

2.64

2.20

2.27

3.01

3.32

1.10

3.07

1.02

2.24

0.74

3.10

1.97

0.64

2.39

0.77

1.86

0.60

3,74

3.15

0.84

2.72

0.73

1.84

0.49

5.73

2.51

0.44

3.01

0.52

1.94

0.34

4.33

4.28

0.99

3.48

0.80

1.95

0.45

AVERAGE: 3.46

3.03

1.01

3.28

1.12

2.10

0.75

Klq : COMPUTED VALUE

- 146 -

■o

FIGURE 3

3 4 5 6 7 8
OBSERVED Kl (m/day)

- COMPARISON OF K|_ FROM EQUATION 21
WITH THE OBSERVED VALUES

- 147 -

FIGURE 4

4 5 6 7
OBSERVED Kl (m/day)

- COMPARISON OF Kl FROM EQUATION 22
WITH THE OBSERVED VALUES

- 148 -

values obtained from Klq = 0.753 <!_. Herein these Klq values are
referred to as the "observed" values, whereas the values, Klc, obtained from
Equation 19 are termed the computed values for differentiation. The predictions of
Equation 22, given in Table 5, are seen to range from 1.87 to 10.10 m/day, whereas
the "observed" values are in the range 0.46 - 7.66 m/day. The average of the
"observed" and computed K|_ values are 2.00 and 4.69 m/day, respectively and the
ratios between the computed and "observed" values range from 0.75 to 7.79, the
average ratio being 3.53. These results show that Equation 22 overestimtes Kl
values for the streams and rivers for which data are reported by Rathbun and Tai
(16).

The predictive capability of Equation 12 developed by Rathbun and Tai (17) is also
evaluated by utilizing the data for the streams and rivers given in Table 1. The
computed values, Klq, for ethylene gas, obtained from Equation 12, are
presented in Table 4, the values being in the range 0.74 - 3.44 m/day. The average
values of Kl and Klc ^"^^ 2.1 and 3.46 m/day, respectively. The Klc
values are 0.34 to 2.27 times the observed values with an average ratio of 0.75
(see Table 4). Thus, Equation 12 is found to underestimate Kl values of
ethylene gas for the streams and rivers listed in Table 1. This is consistent with
the above finding that Equation 22 overestimates the Kl values for the streams
and rivers for which data are reported by Rathbun and Tai (16). These
discrepancies are likely to be due to the possible effects of such factors as wind
velocity, bed slope, presence of other chemical compounds in the streamwaters and
the indirect method of estimating the "observed" Kl values for ethylene.

Sensitivity Analysis

A knowledge of the sensitivity of various dependent parameters on the independent
variables appearing in Equations 19 - 22 will aid in identifying the relative
importance of the dependent parameters In estimating Kl- Herein, the
sensitivity analyses are carried out by the method of determining the relative
error or relative change (described in standard text books on calculus). Equations
19 - 22 are of the general form

Y = ao \^' A' X^n (23)

- 149 -

TABLE 5

COMPARISON OF PREDICTIONS OF EQUATION -22
WITH DATA OF RATHBUN AND TAI (16)

(m)

(m/s) (

m/day)

(m/day)

«L0

0.472

0.160

2.31

3.91

1.69

0.285

0.305

2.51

4.88

1.94

0.289

0.439

7.66

5.78

0.75

0.387

0.317

2.47

5.20

2.11

0.053

0.160

2.12

2.82

1.33

0.090

0.150

1.46

2.96

2.03

0.082

0.150

1.33

2.92

2.20

0.070

0.060

0.46

1.87

4.03

0.260

0.160

0.86

3.58

4.08

0.270

0.085

0.48

2.69

5.63

0.180

0.210

0.92

3.84

4.19

0.280

0.180

1.73

3.82

2.21

0.482

0.160

1.04

3.92

3.77

0.290

0.270

1.39

4.63

3.32

0.518

0.408

0.78

6.10

7.79

0.579

0.552

5.79

7.13

1.23

0.270

0.088

0.68

2.73

4.04

0.948

0.144

1.25

4.14

3.32

0.701

0.171

4.78

4.28

0.90

1.240

0.546

0.81

7.95

9.82

1,270

0.539

1.46

7.94

5.43

2.330

0.747

1.73

10.10

5.85

AVERAGE :

2.00

4.69

3.53

K- : "OBSERVED"

i K,

: COMPUTED

- 150 -

in which Y is an independent variable; Xj (i=l,n), are the dependent
variables; n is the number of dependent variables; ao is a constant; and
ai Is the exponent of X^ (i=l,n). In order to determine the relative
change in Y due to a given relative change in X^, Equation 23 is differentiated
with respect to Xi and rearranged to obtain

dY = ai dXi (24)

Herein, (dXj^/Xi) and (dY/Y) denote relative changes in Xi and Y,
respectively. Thus, for a given relative change in X^, the relative change in
Y can be obtained from Equation 24.. Similar analyses can be carried out for the
other variables X] {i=2,n) appearing in Equation 24. Using this approach, the
relative change in the independent variable due to a relative change of 20% in the
dependent variable under consideration, has been determined. The results are
summarized in Table 6.

Evaluation of Relationships
For Adjustment of K| Values

Laboratory experimental data published in the literature (6, 13, 16, 19, 21) are
utilized to evaluate the validity of Equations 10 and 11.

Rathbun and Tai (16) present K(_ values for oxygen and four organic compounds
whereas Rathbun et al (19) report desorption - absorption rates (per day)
associated with ethylene, propane and oxygen. The studies of Roberts and Oandliker
(21), and Kaczmar, et al (6) include absorption rates of oxygen (per day) and
desorption rates (termed mass transfer rate constants expressed in the units per
second) for several volatile compounds. Thus, exchange rates for oxygen are
measured in all of these studies. The analyses presented below involve ratios of
exchange rates between oxygen (or other tracer gas) and an organic compound. These
ratios have been determined by using either the Kl values or the absorption and
desorption rates. For the sake of convenience, the exchange rates (i.e., liquid
film coefficients, and absorption and desorption rates) will be denoted by K'.

The following relationships are utilized to compute the ratios of exchange
coefficients:

- 151 -

TABLE 6
RESULTS OF SENSITIVITY ANALYSES

Dependent

Variable

Independent Variable

Relationship

Paraneter

Relative Change

Parameter

Equation 19

U2/0ffl

20. Q

-37.8

UZ/v

20.0

37.0

B/Z

20.0

8.0

Kl/U

uVgZ

20.0

7.0

LS/Z

20.0

3.2

Equation 20

UZ/Dm

20.0

-31.8

UZ/v

20.0

30.2

Kl/U

uVgZ

20.0

- 4.6

Equation 21

20.0

37.8

20.0

-37.0

20.0

8.0

20.0

5.2

20.0

- 5.0

20.0

3.2

Equation 22

20.0

31.8

20.0

-30.2

20.0

9.2

20.0

3.0

- 152 -

Re = (K' for oxygen/K' for ethylene) (25)

Ro2 = (K' for oxygen/K' for organic compound) (26)

f^ORG ~ ('^02/Re) - C^' ^01" ethylene/K' for organic compound) (27)

The ratios obtained from Equations 25 - 27 will be termed the observed ratio values
since they are obtained from measured K' values. Table 7 shows the observed
R02 and Rqrq values for various volatile compounds that were included in
various studies (5, 16, 19, 21). The Henry's law constants and molecular weights
of the compounds are also given in this table. The compounds with
H > 10"' atm-m^/mol are listed separately in Table i ,

Mackay and Yuen (13) obtained liquid film coefficients (<[_) at six wind speeds
(range 5.96 - 13.2 m/s) for eleven organic compounds (see Table 7). Since oxygen
has not been included In this study, the values of RoRG» defined by Equation 27
were calculated indirectly. First, the ratios, Rgz* defined by the ratio (K'
for benzene/K' for organic compound), were computed by considering benzene as the
reference compound. Then, these ratios were converted to the Rqrq values
through the use of the observed ratio (K' for ethylene/K' for benzene) = 1.29. The
values of Rqz utilized herein are the average values of the ratios obtained for
each of the windspeeds. Benzene and toluene were included in all the six windspeed
experiments, wheras the remaining compounds were monitored at three or four
windspeeds. The observed ratios, Rqrg* ^re presented in Table 7.

Four of the compounds listed in Table 7, namely benzene, toluene, chloroform and
carbon tetrachloride, were utilized in two or more studies. The observed values of
RqRG ^0^ ^^^^ compound, obtained from these studies, are seen to differ
somewhat, indicating the possible effects of testing conditions on the observed
RqrG values.

The critical volumes, Vq, for various volatile compounds were determined from
the Lydersen method (9). The values of V^ and molecular weights, M, for
various compounds given in Table 7 were utilized to calculate the values of Ry
and Rm according to Equations 10 and 11, respectively. The computed values of
Rv and R^ for various compounds are also given in Table 7. The average
values and standard deviations of the observed and computed ratios are presented in
Table 8. A comparison of the computed Ry and Rm values with the

TABLE 7

RATIOS OF EXCWWGE COEFFICIENTS FOR VOtATlLE COHPOUNOS

Source

Data

Co^XHind

Henry's Lew Constant
(itn-aVnol at 20°C
and I atn)

Molecular
Weight
(gram/ml )

Crttlcal
VoluM
(c»V«ol)

Observed Ratio

Ro?

"org

Equation
Rv 02

Cowxited Batio

To Equation i i

"VORG RmoZ RmORG

Henry's Law Constant
> 10'* at»-ii'/aol

Rathbun et a1

Oxygen

0.73

32.il

73.4

1. 00

l.OO

1.00

(19)^

Ethylene

> 8.6

28.1

120.0

1.15

1.00

1.17

1.00

0.943

1.00

Propane

1.6 X 10'*

44 a

205.0

1.39

1.21

1.38

1.18

1,17

1.24

RatMMrv& Tal
(16)^

Benzene

4.4 X 10"*

78.1

262.0

1.49

t.29

1.49

1.27

1.56

1.65

Toluene

5.2 • 6.6 X

10"*

92.2

316.0

1.56

1.36

1.58

1.35

1.70

1.80

Methylene chloride

3.0 X 10'*

84.9

193.0

1.45

1.26

1.35

1.15

1.63

1.73

Chlorofora

3.4 X 10'*

119.4

238.0

1.50

1.30

1.45

1.24

1.93

2.05

Roberts and

Chlorofoni

3.4 X 10"*

119.4

238.0

1.79

1.56

1.45

1.24

1.93

2.05

Dandliker

Carbon tetrachloride

2.3 - 2.5 X

10''

153.S

277.0

1.62

1.41

1.52

1.30

2,19

2.32

(21)''

1 ch 1 orod 1 f 1 uoroaethane

1.5

120.9

215.0

1.54

1.34

1.40

1.20

1,94

2.06

Tetrachloroethene

8.3 X 10"*

165.8

308.0

1.66

1.44

1.57

1.34

2.28

2.42

1,1,1-trlchloroethane

3.6 - 18 X

10"*

133.4

283.0

1.67

1.45

1.53

1.31

2.04

2.16

Trichloroethene

1.0 X 10 '

13U4

268.0

1.63

1.42

1.50

1.29

2.02

2.14

Kaczaar et a]

Chlorofoni

3.4 X 10'*

114.4

238.0

1.59

1.38

1.45

1.24

1.93

2.05

Broood 1 ch 1 oronet hane

1.3 X 10'*

163^8

259.0

1.82

1.58

1.49

1.27

2.26

2.40

Hackay and Yuen

Benzene

4.4 X 10'*

78.1

262.0

1.29

1.49

1.67

(13)"^

Toluene

5.2 - 6.6 X

10'*

92.2

316.0

1.2S

I. $8

1.80

I,2-d1ch1oropropane

2.1 X 10'*

112.6

299.0

1.33

1.^

2.00

Chlorobenzene

2.6 X 10"'

310.0

1.22

1.35

2.00

Carbon tetrachloride

2.3 - 2.5 X

10"*

153.8

277.0

1.24

1.30

1.75

Henry's Law Constant

< 10 * ata-aVaol

a. I X 10'"

2Q8#3

280.0

2.38

2.07

1.52

1.30

2.55

Kaczaar et al

ChlorodlbroaoKthane

2.70

Hackay and Yuen

Bronofona

5.8 X 10"*

2K.B

301.0

3.45

3.00

1.56

1.33

2.81

2.98

l.?-dlbroMMethane

6.3 X 10"*

m;9

290.0

1.56

1.32

2.49

2-pentanone

3.2 X 10"»

320.0

3.26

1.36

4.75

(13)

2-heptanone

9.0 X 10"*

114.2

430.0

2.77

1.49

2.02

1-pentanol

1,0 X 10*»

88.1

333.0

12.40

t.3tt

t.n

2-Methy1 - l-propano1

1.0 X 10"*

74.1

274.0

14.40

I.M

1.62

1-butanol

5.6 X 10"*

74.1

278.0

21.52

1.30

1.62

Average of the nunber of tests stated below:

' - Bl tests for ethylene and 34 tests for propane
^ - 10 tests for each of the four compounds
- 16 tests for each of the stx compounds

- 13 tests for chloroform and 2 tests for each of the other three compounds

- 6 tests for benzene and toluene, 4 tests for the next three compounds

- 4 tests for the first three and 3 tests for the last three compounds

- 154 ^

Table 8
AVERAGE RATIOS AND STANDARD DEVIATIONS

Henry's Law

Number of
Compounds

Rqrg

RVORG

Rmorg

(atm-m'/mol)

Ave. Std.Dev.

Ave.

Std.Dev.

Ave.

Std.Dev.

> 10"'
< 10"'

18
8

1.35 0.11
7.62 7.50

1.30
1.35

0.10
0.065

1.96
2.12

0.30
0.53

- 155 -

corresponding observed values for various compounds indicates that: (1) the
computed values of Ry. obtained from Equation 10 using the critical volumes,
and their average value, agree reasonably with the corresponding observed values
for the compounds for which the Henry's law constants, H, are greater than
10'^ atm-mVmol; however, for compounds with
H < 10'^ atm-mVmol, the observed and computed Ry values are not in
agreement; and (2) the computed values of Rm. obtained from Equation 11 using
the molecular weights differ from the corresponding observed ratios for about SOX
of the compounds listed in Table 7 (regardless of the H value).

This analysis suggests that Equation 10 can be utilized to compute the values of R
required for adjusting the Kl values of highly volatile compounds (Henry's law
constant >10"^ atm-mVmol). It is recommended to further evaluate the
validity of Equations 10 and U for a broader range of organic compounds as data
become available, as well to undertake investigations for developing similar
relationships for compounds of moderate and low volatility.

APPLICATION

In order to aid in the computation of liquid film coefficient of an organic
compound using the relationships developed in this study, a procedure is presented
below and illustrated by an example.

Computational Procedure:

1. For the given organic compound, obtain Henry's law constant (Note: This
computational procedure Is generally valid for compounds having Henry's
constants greater than 10~ atm - m'/mol).

2. Obtain hydraulic data for the desired reaches (or segments) of a given river
channel .

3. For the desired river water temperature, obtain the values of density and
viscosity, which are readily available in standard fluid mechanics text books
or handbooks on physical and properties of compounds.

- 156 -

4. Determine the molar volume, Vb, for the tracer gas by the LeBas method
(9). For ethylene gas, Vb = 44.4 cmVmol.

5. Calculate Dm for ethylene gas by using the Hayduk-Laudie relationship

(Equation 18).

6. Compute the liquid film coefficient (Kle) for ethylene gas from
Equation 21 or 22, depending on whether the bed slope is known or not.

7. Calculate the critical volume (cmVmol) of the organic compound by
Lydersen's method (9).

8. Determine the ratio (Ry) required for calculating Klqrg from K|_e,

using Equation 10. (Note: Critical volume for ethylene gas =120
cm Vino 1 .)

9. Calculate KlqRG fo'" ^^^ organic compound from

Klqrg = kle/Rv

Example : The hydraulic data for the Speed River, survey date 8/24/78,
reach BD, given in Table 1, have been considered in this illustrative example,
The organic compound selected is methylene chloride (CH2 CI?)- The
computations as per the above step-by-step procedure are as follows:

1. For methylene chloride, Henry's law constant = 3 x 10

atm-m^/mol.

2. The hydraulic data for the reach BD of the Speed River are as follows:

Q = 2.43 mVs; L = 2380 m; S = 0.0011; 8 = 38.6 m;

2 = 0.414 m; U = 13.392.0 m/day; LS = (2380) (0.0011) = 2.618 m

- 157 -

3. River water temperature = IQ.O^C

p = 998.4052 kg/m^; u = 88.73 kg/m-day = 1,027 centipoise
V = u/P = 0.0889 mVday

4. The Molar volume, Vg. of ethylene gas (C2H4) by the LeSas

method is 44.4 cmVmol

5. From Equation 18,

Dm = 13.26 X 10"' = 1.3773 x 10** cmVs

= (1.3773 X 10*M(cmVs)(86400)(s/day)yr(10'*)(cm2/m^3
= 1.19 X lO"** mVday

/day

6. From Equation 21,

KlE =[(2.7) (7.3206 x lO^**)**''* (0.0889)"^-^=' (1.19 x lO"**)^'"'
(38.6)°-'* (0.414)"°-^5 (13392)&.2« (2.618)°-^*']
=4.28 m/day

7. The critical colume, V^, of methylene chloride (CH2CI2) 1

calculated by Lydersen's method is 193 cm^/mol. The critical volume
of ethylene gas is 120 cm^/mol.

8. From Equation 10,

Rv - (193/120)""^^'* = 1.16

9. The liquid film coefficient of methylene chloride is

Kl = (4.28/1.16) = 3.69 m/day
ORG

- I5i

SUMMARY AND CONCLUSIONS

Volatilization of organic compounds in streams and rivers has been investigated
through an analysis of field data on tracers and development of predictive
relationships. The liquid film coefficients of ethylene gas have been determined
by making use of data on ethylene and rhodamine WT dye tracers which were collected
in five shallow streams and rivers located in Ontario, Canada. The development of
generalized relationships among the liquid film coefficients and bulk flow channel
characteristics has been accomplished through dimensional analysis techniques. The
predictive capabilities of the relationships have been evaluated by a comparative
study of the observed and the predicted values of Kl for ethylene gas
applicable to the five Ontario streams as well as other streams for which data are
reported by Rathbun and Tai (16). A sensitivity analysis of various parameters
appearing in the relationships has been carried out. The relation of liquid film
coefficient of an organic compound to the molecular weight as well as critical
volume (or molecular diameter) of the compound have been evaluated with the aid of
laboratory experimental data reported in the literature. The computation of liquid
film coefficient of an organic compound using the relationships developed herein is
outlined in a step-by-step procedure, and illustrated by an example.

The conclusions of this investigation are as follows:

1. The predictions of the relationships among liquid film coefficient and bulk
flow characteristics derived from dimensional analysis procedures compare
satisfactorily with the observations.

2. Application of Equation 22 developed from the dimensional analysis procedure
to the streams for which data are reported by Rathbun and Tai (16) results in
an overestimation of the liquid film coefficients for ethylene. The relation
obtained by Rathbun and Tai (17) is found to underestimate the liquid film
coefficients of ethylene for the five streams and rivers considered herein.

3. The ratios (Kl for ethylene gas/Kt_ for organic compound) obtained from
experimental data are in reasonable agreement with the values computed by
using the critical volumes of the compounds (Equation 10) but differ from
those obtained by using molecular weights (Equation 11) for organic compounds
having Henry's law constants > 10"^ atm - mVmol. Thus, the

- 159 -

ratios obtained from Equation 10 are applicable to calculate Kl values of
highly volatile organic compounds from those of tracer gases.

The results of this study indicate that the relationships developed herein (i.e..
Equations 21 and 22) and those of Rathbun and Tai (16, 17) may not be universally
applicable to streams and rivers of differing hydraulic and water quality
characteristics, as well as for different organic chemicals of public health
concern. In general, there is a lack of data required for development and
validation of generalized relationships among Kl, channel hydraulic parameters
and properties of chemicals. In light of these deficiencies, the following areas
for further work are recommended:

1. The effect of wind velocity on Kl under field conditions should be
investigated and development of modified relationships for Kl should be

undertaken.

2, Investigation of relationships between Kl and molecular properties for
chemical compounds with Henry's constants lower than
10"^ atm - m^/mol (i.e., moderate and low volatile compounds)

is recommended.

- 160 -
APPENDIX - REFERENCES

1. Bacchus, A. "Field Measurement of Stream Reaeration Coefficient", Water
Resources Paper 13, Water Resources Branch, Ontario Ministry of the
Environment, Toronto, 1981, 41 pp.

2. Cohen, Y., Cocchio, W, and Mackay, 0. "Laboratory Study of Liquid-Phase
Controlled Volatilization Rates in Presence of Wind Waves", Environmental
Science & Technology, Vol. 12, 1978, pp. 553-558..

3. Gore & Storrie Ltd., "Volatilization Rates for Organic Chemicals of Public
Health Concern", Technical Report prepared for the Ontario Ministry of the
Environment, Toronto, Canada, 72 pp.

4. Dilling. W.L. "Interphase Transfer Processes, II. Evaporation Rates of
Chloro Methanes, Ethanes, Ethylenes, Propanes, and Propylenes from Dilute
Aqueous Solutions, Comparison with Theoretical Predictions", Environmental
Science & Technology. Vol. 11, No. 4, 1977, pp. 405-409.

5. Hayduk, W. and Laudie, H. "Prediction of Diffusion Coefficients For
Non-electrolysis in Dilute Aqueous Solutions", American Institute of
Chemical Engineering Journal, Vol 20, 1974, pp. 611-615.

6. Kaczmar, S.W., D'ltri, P.M. and Zabik, M.J. "Volatilization Rates of
Selected Haloforms from Aqueous Environments", Short Communication.
Environmental Toxicology and Chemistry, Vol. 3, 1984, pp. 31-35,

7. Lewis, W.K. and Whitman, W,G, "Principles of Gas Absorption", Industrial
and Engineering Chemistry, Vol , 16, No. 12, 1924, pp. 1215-1220.

8. Liss, P.L,, and Slater, P.G. "Flux of Gases Across the Air-Sea Interface",
Nature. Vol. 247. June 25. 1974, p. 181-184.

9. Lyman, W.J., Reehl, W.F, and Rosenblatt, D,H. (eds) Handbook of
Chemical Property Estimation Methods - Environmental Behaviour of
Organic Compounds , McGraw-Hill Book Co., Toronto, 1982.

10. Mackay, D. and Wolkoff, A.W. "Rate of Evaporation of Low Solubility
Contaminants from Water Bodies to Atmosphere", Environmental Science &
Technology, Vol. 7, 1973, pp. 611-614.

11. Mackay, D, and Leinonen, P.J. "Rate of Evaporation of Low Solubility
Contaminants from Water Bodies to Atmosphere", Environmental Science &
Technology, Vol. 9, 1975, pp. 1178-1180..

12. Mackay. D. and Yuen, T.K. "Volatilization Rates of Organic Contaminants
from Rivers". Water Pollution Research Journal of Canada. Vol. 15. No. 1,
1980, pp. 83-98.

13. Mackay. D. and Yuen. T.K. "Mass Transfer Coefficient Correlations for
Volatil ization of Organic Solutes from Water", Environmental Sciences
Technology. Vol. 17. No. 4, 1983. pp. 211-217.

- 161 -

14. Perry, R.H. and Chilton, C.H. (eds). Chemical Engineers' Handbook , 5th
edition, McGraw-Hill Book Co., New York, 1973.

15. Rainwater, K.A. and Holley, E.R. "Laboratory Studies on Hydrocarbon Tracer
Gases", Journal of the Environmental Engineering, ASCE, Vol. 110, No. 1,
1984, pp. 27-41.

16. Rathbun, R.E. and Tai, D.Y. "Technique for Determining the Volatilization
Coefficients of Priority Pollutants in Streams", Water Research, Vol. 15,
1981, pp. 243-250.

17. Rathbun, R.E. and Tai, D.Y. "Volatilization of Organic Compounds from
Streams", Journal of the Environmental Engineering Division, ASCE,
Vol. 108, No. EE5, 1982, pp. 973-989.

18. Rathbun, R.E., Shultz, D.J., and Stephens, D.W. "Preliminary Experiments
with a Modified Tracer Technique for Measuring Stream Reaeratlon
Coefficients", United States Department of the Interior Geological Survey
Open File Report No. 75-256, Bay St. Louis, Mississippi, 1975.

19. Rathbun, R.E., Stevens, D.W., Shultz, D.J., and Tai, D.Y. "Laboratory
Studies of Gas Tracers for Reaeratlon", Journal of Environmental
Engineering Division, ASCE, Vol. 104, No. EE2, 1978, pp. 215-229.

20. Reid, R.C., Prausnitz, J.M. and Sherwood T.K. The Properties of Gases
and Liquids , 3rd ed., McGraw-Hill Book Co., New York, 1977.

21. Roberts, P.V. and Dandllker, P.G. "Mass Transfer of Volatile Organic
Contaminants from Aqueous Solution to the Atmosphere during Surface
Aeration", Environmental Science & Technology, Vol. 17, No. 8, 1983,
pp. 484-489.

22. Rouse, H. Fluid Mechanics for Hydraulic Engineers . Dover Publications
Inc., New York, 1961.

23. Smith, J.H. and Bomberger, D.C. "Prediction of Volatilization Rates of
Chemicals In Water", Water: 1978 AIchE Symposium Series, Vol. 190, 1978,
pp. 375-381.

24. Smith, J.H., Bomberger, D.C. and Haynes, D.S. "Prediction of the
Volatilization Rates of High Volatility Chemicals from Natural Water
Bodies", Environmental Science & Technology, Vol. 14, 1980, pp. 1332-1337.

25. Southworth, G.R. "The Role of Volatilization In Removing Polycyclic
Aromatic Hydrocarbons from Aquatic Environments", Bulletin of Environmental
Contaminants and Toxicology, Vol. 21, 1979, pp. 507-514.

26. Weast, R.C. and Astle, M.J. (eds). Handbook of Chemistry and Physics ,
60th ed., CRC Press, Inc., West Palm Beach, Florida, l9«u-al.

- 162 -

ACKNOWLEDGEMENTS

The project was supported by a research grant from the Research Co-Ordination
Office, Ontario Ministry of the Environment, Toronto.

Constructive comments offered during the study by Messrs. Dennis W. Draper and John
G. Ralston, Water Resources Branch, Ontario Ministry of the Environment, Toronto,
are gratefully acknowledged. Thanks are due to Mr. Allan F. Bacchus for his help
in making field data available for this study.

- 163 -

VOLATILIZATION RATES FOR ORGANIC CHEMICALS OF PUBLIC HEALTH CONCERN
Technical Report prepared b^^ Gore & Storrie, Ltd., March 1984.

Errata

Page 15 - Table 3.1, column 5: {m /s) instead of (m /s).
Page 24 - Line 16: 3.59 instead 3.59%
Page 25 - Line 5: 6.02 instead of 6.02% q ^6

Page 26 - Line 12: Equation 4.9 should include U ' .The corrected
equation is as followsi

Kl = 2.70 g °'^* V "'-'^ Dm '''^ B"'"* Z"°-** U"-^^ (LS)"-^^ (4.9)

Page 60 - The following reference should be added:

Liss, P.S. and Slater, P.G., 1974. Flux of Gases Across the Air-Sea Interface
Nature, 247: 181-184.

- 165 -

Experimental and Environmental Modelling Studies

of
Hazardous Chemicals

D. Mack ay

S, Paterson

n. Cheunji

W.Y, Shiu

Department of Chemical Engineering and Applied Chemistry
University of Toronto

- 166 -

ABSTRACT

I'rot.ress is described on an environmental modelling project whicli, it
ir. hoped, will be used to assess the behavior of chemicals in Ontario, A
Level III f'u(^acity nocel has been conipiled which can be used to calculate the
behavior of a chenical which is subject to steady state partitioning, reaction,
advection and interconpartnental transfer In an evaluative environnent
consistinj^ of six conpartnents, air, soil, water, bottom and suspended
sedinents antJ fish. The nodel is applied to 14 chemicals of varying properties
and is stiowr to generate behavior profiles which are consistent with reported
cheiiiccil fete observations in the real environnent. Since this single model
;;enorates a set of consistent behavior profiles, it is suggested that it nay be
useful for predicting the behavior of chemicals for which no environmental
observations yet exist. Environmental processes which are still inadequately
treated by the model sre discussed with a view to later improving the model's
predictive reliability.

In a [parallel effort (published in July 198^) we have addressed the
issue of characterizing the heterogeneous spatial distribution of a chenical
using probability density functions.

A 'Southern Ontario model is being developed which attempts to combine
the principal features of the evaluative model with the spatial distribution
approach and can be used to assess the environmental fate of chemicals in that
region.

- 167 -

INTRODUCTION

In 1912 we started a tliree-year project to develop a couifHjter model
which could be used to help assess the behaviour of existing and new
environr,ental chc-mical contaninants in Ontario. In this paper we review
prof.ress in this project. Cone details of the justification for the use of
nodels and the- approach bein?;, adopted were civen in a previous Technology
Transfer Proceedings (tlackay et al 19£3).

V.v have elected to develop two nodels. The first is a purely
evaluative nodel in which a hypothetical environment is assenbled corsistinp of
£jn arou of 1 square kilorietre containinr^ reasonable volunes of soil, air,
water, sedinont (both botton and suspended) and fish. Typical interphase icg.
dir-Water) transfer rates are assigned and the behavior of the chemical in
that evaluative environment is computed. From this a "behavior profile" is
ot:t;iined ,

The second is an adaptation of the model to have volumes and areas
which are representative of Southern Ontario and which can be used to predict
the l..eliavior of cheniicals in the Province, provided that data .ire available
for emission rates and the chemical's partitionini^, reaction, and transport
propertie.s.

ft third effort has involved the establishment of a procedure for
partially "validating," the Southern Ontario model by comparing computed
concentrations with actual monitoring data. Since concentrations vary in tine
and space it is necessary to characterize these variations by a statistically
riLorous procedure. Development of this procedure has received special study.

- 168 -

A fourth effor-t has involved a number of auxiliary studies such as
'iLt(?rr;ination and corrrlation of physical chemical properties for the
substances nf interest.

EVALUATIVE MODEL

A ro[ ort has been coiipleted on this aspect of the work and has been
sulHiitted for publication to Chemosphere (Mackay et al 1584). V.o benefitted
(greatly by the voluntary association of Dr. V., Drock Neely of Tow Cheinical Co,
llidland, t'T who is an acknowledf.ed expert in this area.

A Level III fuf;acity :nodel has been developed containing six
conpartnents (as shiown in the Figures) and for which expressions are derived
ctjbrjcterizin.: partitioning, reaction and transport properties. The model has
been applied to 1M chemicals of varying properties and has been "fine-tuned" to
e.eneratG huhavior profiles which are consistent with observed behavior. The
chenicals and their properties are listed in Table 1. The output which takes
the forn of a computer output nunerical listin;'. is converted by hand into "fate
diagrams" as illustrated in Fit?,ures 1 to 3. These diagrams contain the
essontiul Mass balance information in a condensed for-ri.

It is encournginr that this one model is capable of treating, chemicals
whif h dift':r so r.^eatly in p.hysicol clienical properties, reactivity and
tr;:tisport characteristics. It is believed that the model will he useful for
pr>!uictir;g in advance the behaviour of new chencials or chenicals for which
there is insufficient environmental experience. Me emphasise that our aim has
l)et-t, to i;rx'sc-nt and justify the tnethodology rather than give definitive
choMical fate asscssnents. It is expected that users will select different,

Coopound

Molecular
Weight

DDT 354.5

Mlrtx 545.6

TrlehlorMtlqrl«n* 131.4

r«nicrothloo 277.0

AtraslM 215.7

Aldicarb 190.3

Chlorprrlfoa 350.6

Aiichr«c«fM 178.2

HonochlorolMnssfM 112*6

1.4-Dlchlorobcns«M 147.0
1.2.4-Trlchlorob«itt«M 161.5

Htxachlor obcnseiM 284 . 8

BcnssiM 78. 1

p-cr«sol 108. 1

Table 1

Phyelcochealcel ProperCiee

Vepour

Solubility S

Octanol-weter (logK^)

5
Pressure F (Ps)

(g/«')

Partition Coefficient

1.33E-OS

1.70E-03

5.98

1.33E-04

7.0E-05

6.89

9.87E-^03

l.lE-f03

2.29

9. 19E-03

2.7E'H)1

2.33

3.0E-06

3.3E+01

2.33

1.33E-02

6.0E-H)3

0.70

2.53B-03

2.0E400

4.99

2.3E-05

7.3E-02

4.45

1.57E^3

4.88B-K)2

2.86

4.75E-IOI

7.0E-H)1

3.42

6.08E>01

2.5E+01

4.04

3. lOE-03

6.0E-03

5.61

1.27E^4

1.78E403

2.13

1.44B401

1.80B403

2.20

- 170 -

Mori ii[:\>ro',)r iiilv cor; pari. iiicnl, volunec and areas, and that sone of the constants
in the nodcl such ns the sedinient-v/ater, soil-air and soil-\jater transport
Fcirur.etcr:; nay require nodif ication. They appear however to be of the correct
order of nai^nitude.

It is expected that other conparttnents and processes will be included
in the future, notably wet and dry deposition of atmospheric particulates and
leachiHi^ into ^.round water. This can be accomplished with little increase in
proiirarn complexity.

i;e suLXest that reliable assessment of chenical fate requires the
consideration of partitionin^i, reaction, and transport data in sone form of
evaluative model such as fut;acity Level III. It is rarely possible to process
or assimilate the data by other r.eans. The chemicals considered here
illustrate tfie vjide diversity of behavior which is likely to be encountered.
Pictorial representations of the output are believed to be useful for
coMMunicatin;., the results. An attractive feature of the model is that it can
be used as a starting point for determination of the dominant conpartnents and
processes of interest. It nay then be appropriate to examine these
con;parti'ients and processes usinf, another model, more limited in scope, but more
accurate in its detail, and possibly site specific.

A "uscr-fricndly" version of the model has been compiled and is being
made available oti request on a cost-recovery basis. A disk is available for
use on the IB!! PC systems and one v.'ill be available shortly for the Apple
systen, Tfie ine[,.ory usa^;e is minimal.

- 171 -

An aspect of this work which is receiving-, particular current attention
i.s ii sonsitivity analysis, i.e. how sensitive the output is to variations in
ttio input parancters,

COKCENTRATIOK DISTRIBUTIONS

'./hen coriparin:; the nodel predictions witli field data an irir.iediate
problen arises in that one-to-one comparison is not possible. Environmental
concentrations vary in tine and space thus we are conparin^; one value with a
distribution. An alternative approach is to devise nethods by which the model
predicts not a single value but a distribution which can then be compared with
the actual distribution.

A paper on this topic has been prepared and published in Envirotlrichtal
Science and Technolo£y in July 19fi^ (Hackay & Paterson VJci'J). V/e favour the
ULC of l.cibull distributions rather than lognorrial distributions for this
purpose because of their greater flexibility and mathematical convenience.
Fit^ure '1, which is taken from that paper, illustrates the two processes of:-

a) taking real environrental concentration data and converting, then into
liistoj.rai;is and cumulative distribution functions, from which means and "spread
factors" can be deduced,

b) The reverse process of takinr-, evaluative environmental concentrations or
amounts and assigning then to distributions on the basis of experience with
sit.ii lar chenicals .

- 172 -

Thir. ultir.utely can loud to distributions es shown in Tinure [; of
ccnc'^ritrati-:)r;s in various nedia. The proxinity of these concentrations to
actu.il "cffoct levels" can then be exai.iined. fore details is j^aven in the
paper.

SOUTHERH OriTARIO nODEL

71ie evaluative model and the conpontration distribution work
coalesce in the Southern Ontario nodcl.

This nocel is specifically desit;ried for investicating chemical
beltcvior in the .Southern Ontario environment. Its structure is almost
identical to that of the evaluative r.odel as described in the paper submitted
to Chetno-phere (r'ackay et al IQf^^'O except as follows:

1. The :^tudy ar.-a is increased and the relative compartnental volumes are
nodified to sinulate Southern Ontario,

2. The ei ission rate is set specifically for each chemical using available
enission rate data.

3. The advection rates in p.ir and water conpartnents are set for Southern
Ontario.

M. The air-water riass transfer coefficients. K-,2 and K^^, are calculated from
nass transfer correlations p.iven in tiackay and Yeun (1983). The air and
water diffusivities, Dt;^ and 'J\■\^ are estimated usinf. the method of Fuller,
lichettler and Giddincs (1966) and l.'ilke-Chani'. (19!}5) respectively.

STUDY AREA " "^^^ ~

ritj,ijre 6 is a riap sl-iowinc the study ares of Southern Ontario. The
houndijries are drnun to include most of the rer.i'^ns of substantial emission and
thus of contamination concern. The French River was chosen to be the northern
boundary. All of Lake Ontario and Lake Erie north of the international
boundary fori,; the southern boundary. The west side is bounded by port of Lake
liuron including Georgian E^ay and part of Lake St. Clair. Other lakes included
in this retiion are Lake Sincoe and the other inland lakes.

In the interests of simplicity the study environment is acain divided
into G conpartnents. The total area of land is estimated by neasuring r.ap
areas to be 12.^ x lo''^ n^ and that of water is 6,37 x lo'' n . The
calculation of conp^rtrient volune is nore corplex as it involves using either
estimated depth of part of or the whole lake or literature values. The present
estinatep are ;-\iven in Table 2,

TABLE 2

Estinatec areas and volumes of compartnents in Southern Ontario

Coi:;partrient [iorizontal Are^i m V'^lu^'e £1

Air - _ - 3.73 X 10^^

l.'ater 6.37 x ic"*" 3.^6 x 10^^

Soil 12. i* X 10^^ 1.86 X 10^"

Pediment 5,37 x 10^^' 1.91 x 10^

r.uspended Sedinent _ - _ 1,73 x 10'

Aquatic Biota - - - 3.^6 x 10^'

- 174 -

RESULTS

.'i tntiil of twenty-four c^lel;licols which include strazinc,
trichloroethylene, benzene, tv/elve chlorobenzenes, phenol and seven
chlorophenols were subjected to Level III calculations usinj: the Southern
Ontario rriodel, Specimen emission rate and the receiving conpartnent for sone
cheniculs ijrc piven in Table 3. Due to the dearth of enission data, a great

deal of effort was devoted to estimatinc emission rates based on world
[poouction data, nonitoring data of residues or intrapolation between known
enission values of isoners in a liomolosous series. Thus the values given in
Table s are speculative,

TABLE 3

F.r.iijsion rate and receivinc copartnent in Southern Ontario envlronnent
for selected chenicals.

Compound l-nission Hate Receivlnb Conpartment

Cnol/h)

nirpx 2,^32 x 10"^ v;ater

atrszine 5. PC x 1o' water (lOr,), soil (90^)

trichlorocthylone ^.31^3 x 10^ air (90?), water (57), soil C^t)

Ideally, it is desirable to investlc^to the functional dependence of
advection rate on time, season and location. This implies that neteoroloEical
and linnoloi.ical factors be considered, Ey incorporatin-^ these factors in
detail the iiodel will hecone too conplex and tine - and location-specific. It
was thus docideu to use j ami 100 days as residence tine in air and water
conpartrients respectively. The corresponding advection flowrates in air and
water are 3-12 x lu^"^' ir.^/h and l.^iJ x 10^ n^/h.

- 175 -

The noricl results yield concentrations of the chemical in the six
cor pcirtments which can he compared with environnental values as illustrated in
ri^',urc'.'i 7 to 9, Concentration is expressed on a loc scale to accomodate the wide
ranf.e ol' possible values. The environmental concentrations usually exhibit
d rarii',c of vnlu*;s which are best represented hy a bar rather than a single
point. Since the model is capable of eenerating a point value of compartnental
concentration only, the value is depicted as a point in the figures. The
volidity of the Southern Ontario model In predictinp. environnental
concentrations in different conpartments is indicated by the proximity of the
model results to the environmentiil monitoring data.

Fit'.ure 7 is a comparison between model results and environmental
concentrations for r.ilrex. The concentation in water i/as predicted to be low
and a^^reet; with the observed value which is below the detection limit of
C.I uf:/r:-^. The computed value of 0,22 ug/kr. in sediment phase is comparable to
l.f; u^ykh i.s dettrmined from the sediinent data (Thonas and Frank 19^3) by
Mackay and Paterson (19R^). The concentrations in suspended sediment and biota
i;ere predicted to fiUl below the observed values. The sediment phase was found
to be the principal sink or conpartnent of accumulation for mirex. The loading
of mirf.'X in sediment is C3C kt; which is in [iood atreenent with the reported
value of about 700 kc (Ilalfon 198«) based on a ^0t^ survey.

I'onitorint^ cJata of atrazine are relatively rare. The only reported
value is 1000 U{;/m^ which is based on samples taken form streams entering Lake
Ontario CFiLure £). The model predicted the concentration in water to be
;?7 ui/in^. The two values differ by about a factor of forty However when the
dilution effect of the lar[;e volume of lake body ib taken into consideration.

- 176 -

and when it is noted that tho measurenents were taken in areas of atrazlne use,
the difference in Magnitude of concentration is understandable. As the model
shows that utrazine residues mostly accumulate in soil, it would be desirable
to Monitor the herbicide concentration in soil so that data are made available
for Model cor.iparisoti,

Tiichloroethyletie was predicted to have a concentration of 3.7 ul./m^ in
water which falls within the rang-.e of observed values (Figure 9), Despite the
hicJi tendency of trichloroethy lene to partition into air, its concentration in
the air coripartrtent has not been monitored, or data are not available.

Although benzene and nonochlorobenzene are ubiquitous in nature, their
concentrations in various compartments have not been recorded. Consequently
the results fenorated by the model can only be viewed as values of likely
r-iagnitudo sinilar to environr;ental concentrations.

The horolo^oua series of chlorobenzeties was predicted to exhibit
j.enerally sinilar behavior profiles and the concentrations in the respective
coMpurti.Mjnt follow a similar trend. As the degree of chlorination increases
the concentrations of the isonier in air and water decrease; and the
coticentrLjtions in sorted [ifiases increase. Generally, the predicted
concentrations fall within the ranpe of observed values in water, sediment and
bio!,a con} artnefits respectively. In sorio cases, the model overestimated the
concentration in \;3ter (1,3-000, 1,?,3,-TC3, 1,3,5-TCC 1,2,3,5-TeCD) while in
other castJt;, the predicted values are at the low ends of the ran^e of monitored
data (as found in the sediment phase for di- throut'h hexachlorobenzenes). The
.^ood af;recr!ont of concentrations in biota between observed and computed values
is particularly oncouref;inr.

- 177 -

f'onitorinc. data for phenol and chlorophenols are surprisiriiily scarce.
The only data civailable for nodel comparison ore concentrations in water for
?,'l-dichlorophenol (2,J|-DCP) and pentachlorophenol (PCP). These data were
collected fron waste streams discherced from a pulp and paper mill plant (2,4-
ncr) and sev;ai;e treatment facilities in Southern Ontario (PCP) and hence are
expected to be nuch hif.her than those observed in large water bodies such as
the Groat Lakes. This probably explains the deviation of the low
concentrations as generated by the nodel fron the observed values.

In conclusion, the early results are encouraein£ in that most
concentrations are of the correct order of maj^nitude and when large
discrepancies exist they are readily explainable, fiuch more conparison and
interpretation renains to be done before the model can be regarded as being

sufficiently validated.

HUMAN EXPOSURE

One of the end-points of chemical fate calculations is the assessment
of human exposure. The approach of Rosenblatt et al. (1900) has been
tentatively applied to the nodel output to estimate the amount of a chemical
vjhich reaches a human beinf^',. This has been done illustratively for mirex
(Paterson and 'iBckay, Handbook of Envlronriental Chemistry, in press ^9^M), If
an acceptable daily exposure to a toxicant is determined, corresponding
per, lissible concentrations for air, water, crops, anima]s and fish can be
calculated. If these are found to be excessive then measures can be taken to
reduce the emissions accordingly.

- 178 -

The roijt difficult iiodlun to quantify is food. Its extent of
contnnination is difficult to predict and may vary greatly with entry route of
the contar-inant (cc- the chenical nay be directly applied to the food as a
pesticide or may enter the food chain through soil).

For illustrative purposes, and to indicate a future direction for the
incorporation of fate nodels into exposure assessnent, food is defined as
hLtvin^ a fut,acity capacity Z corresponding', to 10" of Z for fish plus lliO'l of Z
for v;ater plus 5f- of Z for soil. It is emphasised that this is purely

r.pecui.'itive ,

Ttiis calculation is applied here to 2,3,7,8 TCDD. By adjusting the
enissions, environmental concentrations in good agreement with literature values
may be ot'tainec (see Table -l). The dominant routes to nan can then be
dcternlned; conparisons can be r.ade with existing or estimated guidelines; and
appropriate rieasurcs can be suggested, if necessary. This is illustrated in
Fi^-.ure 10.

TABLE 4

Predicted and observed environmental concentrations for ?, 3,7,8 TCTD.

Phase

Predicted

air

2 X

10-^^^

water

1.6

X 1.-^^

^oil

2 X

10"''

scdii..i

;nt

<= X

10"^'

biotu

1 X

la-"^

Literature

Ratio Fred, /Lit

3 X 10

-16

2 X 10"''^ ii')

3 X 10"^'

6 X 10

-8

6.7
1/12,5

1/1.5
1.67

- 179 -

Fnr- ;f,l^,Y,!i TCl'D, tho r^ajor nx[osnre route to ncn ir. food, prinarily
fish. Thir. n^iy not bo tho doninnnt dosage route n-s atisorption frorr the <;astro
iritc:3tin,'ii. trijct is not quriPt.i ficd. ilowe^ver, the rnlculatlon provides an
Ddditional dinension to the tehavior profile of the chemical.

COHCLUSIOriS

The wor-k ic- pro[^resEin,. well; ue are encouraf.ed by the success of the
evaluative nodel; the early results of the Southern Ontario are fairly pood and
ttie hunan exposure os;:ect hold:; promise of beint', particularly valuable in toxic
c 1 1 en i c a 1 r i a n a t' e iit n t .

REFERCrXES

!'r,c<ry, D.. Patrrson, S,, Chcuni , F., and r.hiu, \j.y., "Experinental and

[■nvironnental ?'odellinG T-tudies of Hazardous Substances in Ontario.
Proceedinti Technolo;;y Transfer Conference rio. M Part 1 p 37C.
Ontario f'inistry of the Environnent 19r3.

[lac-:ny. P.. Patorson. L\, Cheun^.^, L,, Keely, '.'.P., "fvaluatinr. the

F.nviornnental Behaviour of Chenicals with a Level III Fugacity Model"
fap.er submitted to Cher^osphere T:J3'1.

"ackay, ",, ?aterson, V.,, "Spatial Concentration Pistributions" Environ. Sci.

Tcchnol. 2j_ 2';7A O^SlA)

!:ackr.y, C, Yeun, Aj.r, "toss Transfer Coefficient Correlations for

Volatilization of Organic Solutes froni Water "Environ, Sci. Technol,

Miller, L;.-.., ^chcttler, ?.i'., (biddings, J.C. "A i:cw P.ethod for i inary Gas-
Phase Diffusion Coefficients" Ind. Enp.. Chen. 5H, 19 (1966).

.,ilke, C.l;., Ch.nn^, P. "Correlation of Piffusion Coefficients in Dilute
Solutions" AlCtir: J. J_» 26'! (1^:55).

Thot.as, H.:.., Frarik, R, In "Physical i3ehavior of PCBs in the Great Lakes"

Mackay, D, et al., Eds. Ann Arbor Science: Ann Arbor, lUch., (19P3)

pp 2U:i - C7.

Halfon, E. "Error Analysis and Sinulation of f'lrex Behavior in Lake Ontario"
Fcol. lodelllnL'. 22, ?.^3 C19P3/N).

Paterson, :-'.. and Mackay, D., "The Fugacity Concept in Cnvironnental tlodellirc"
in^ The Fandbook of Rrvironmental Chorlstry (Fd. liutzinj^er) vol 2C,
Springer Verla£ (in press 1904)

'D ■ "t^

- 180 -

Fltur« 1: Envlror»ient«l f«t» dl«gr*in for
Kirex

^ 1.53 X 10"*
/

Flow and reaction « 5,05 y
time

^j^^^*"^ _, _._^ 0.751 Reaction persistence ■ 23.3 vj

1.49x10"

3.20x10 S.Sed.
^ Bed. '

0.0

1-10*

-10

water

S.Sed

.Blotfl

— S«d.

-10""^

Soi]

-10

Water.

Mass distribution

'^Emissions

^" Reaction

^ Transfer
— — —^Advection
(mol/h)

Air

-3

-4

_10

■Air

•-Soil
-10"^

-10

-6

-10

-7

l_10

-8

Removal distribution

Concentrations Fugacity
(mol/m-') (Pa)^

I I R«c

tior
Advection

f = fugacity (Pa) j
C " concentration (tnol/ni )
m • amount (mol )
5 = percentage of total amount

- 181 -

Fiturc 2: Envlron*ent«l f«t» di*cr«r
Atrazlni

^9.97k10" Fioj, ,nd rMCt^on • 0.35 y

/'^ time

^^ 2 28 x lo"

0.0 ^^ \ w Reaction persistence •O-^*^

- 10

Mass distribution

► Emissions

► Reaction

► Transfer

— ■ — ■-^■Advection

Air

-1

i-io

Uio-^

LlO-3

-4

-5

10
.Soil

-7

-8

- 10

Lio-'

■ Sed.

L 10-^'

L io-'2

I- io-'3

1. 10

-u

Removal distribution

I I Reaction
[ I Advectlon

Concentrations Fugacity
(molV) (Pa)

f ■ fugacity (Pa) ,

C - concentration (mol/m )
n « amount (mol )
1 > percentage of total amount

•* Wattr
S.Scd
Biota

- 182 -

Figure 3: Envlronntntal f«t« dl«Rr*Bi for
Trichloroethylene

0.i63

,-^2.07 xlO

Flow and reaction "69.4 h

tlM

_. _._^f, e^^ Reaction persistence "150 h

,A.S1 X 10

~i*

r- 10

-2

-3

0.0

.Soil

HlO^"""
•— Biota

1 Bed.

Ls.Sed.

^°

6.69.

0.0

Biota

0.0

- 10

,3.32 X 10

-7

Biota-

Soil
Sed. ^
S.Sed.J

Uater.

Mass distribution

^' Emissions

^- Reaction

^- Transfer

— •■ ^- Advection

(mol/h)

Air-

-4

- 10

Air
-5

^10-^

-7

_ 10

Lio-a

Removal distribution

j I Reaction

I I Advection

Concentrations Fugaci
(mol/m-'} (Pa)

f = fugacity (Pa) 3

C " concentration (mol/m )

m ■ amount (mol )

' " percentage of total amount

- 183 -

Figure ^. llluscratlon of Forward and Itevaraa Procaaaca for Traaciaent of
Real and Evaluative Concentration Data

FORWARD PROCESS

REAL ENVIRONMENT

ACTUAL
DENSITY FUNCTION

Vt ■ too

CONCENTRATIONS

CUMULATIVE FUNCTION

PWi

t 4 ■

COMCEHTRATION

> 4 fl

COMCNTNATION

NTRATIflM

REVERSE PROCESS

PREDICTEO CONCENTRATIONS

DENSITY FUNCTION

CUMULATIVE FUNCTION

■lih Vf ■ too
Cm ■ 5.ft

■ 4

CONCENTNA1ION

C ai m w Ca • S.B

EVALUATIVE
ENVIRONMENT

••

t 4 «

COHCCNTRATKM

SOURCE

EXCEEOENCE 0.01%

EFFECT

CONCENTRATION

EXCEEOENCE 10%

PERMISSABLE FOOD
CONCENTRATION

EXCEEOENCE 3%

DRINKING WATER
OBJECTIVE

EFFECT
CONCENTRATION

LETHAL
CONCENTRATION

LOG CONCENTRATION

Figure S. Relationship of Predicted Environmental Concentration Distributions to
Toxic Effect Concentrations

FTClIRr 6

FiCURr 6: Southern Ontario Study Ar^a

''""" FIGURE T
COMPARISON BETWEEN COMPUTED • f\ND
OBSERVED □ CONCENTRATIONS FOR MiREX

AIR ■-, 1 ■

/"3

0-* lo^' 10"* 10'^ 10'* x/a/m^

WATER u^_ "^

I 1 1 L

r^ .«-»

10"^ I0-* 10"' lo" 10' /^^/i^

50IL

J 1 1 I

IO-* 10-^ I0-* 10- ,0° A'5/'«5

SEDIMENT i^ 1

J 1 I

'° '0 10' '0> 10' ^3 /kg

SUSP. SED.

J L

»0" 10 10' 10^ 10^ fx^/kg

BIOTA

J I L

lO"' 10** lO'

'^^ 10^ /"5/''^5

"'''" FIGURE 8

COMPARISON BETWEEN COMPUTED • ftNh
OBSERVED D CONCENTRATIONS FOR flTRflZlNE

AIR I .—I I I i

-^5

id"' 10"* 10"* 10'* 10'^ Xiq/m^

WATER I ■ , '. T"" ■

10° lo' 10* 10^ 10* /^S'''^'

SOIL I- _!_, I ■

10' 10' io» 10* itf A«3/'^9

SEDIMENT I i^

O" ICr' 10- 10° 10' /x^/kg

SUSP. SED. ' i •! \ I

0"' 10"' id"' lo' 10' fji^A^

BlOTfl • r '

10"^ 10'^ iQ-' 10° lo' /Jj^q

-188- FIGURE 9
COMPflRlSOhJ BETWEEN COMPUTED • AMD
OBSERVED □ C0^JCENTRflTl0^4S FOR
TRICHLOROETHYLENE

AIR I I »J ■

10"' 10"' 10"' 10° lo' uq/m^

WATER I I I i_

-3 . -2 -I

10 lo' lo' 10° 10 ' /jg/m*

SOI L I •! 1 1 I

icf 10 ' 10 ' 10° lo' H3^^3

SEDIMENT I ..

•O' 10-^ 10- 10° 10' ua/ki

SUSP. SED. ' ^ \ . i

10'^ 10"^ lO"' \(f w' H3^3

eiOTA . I . ■

10'^ 10"^ iC 10° lO' yuj/kg

Figure lO, SPECULATIVE 24-HOUR EXPOSURE OF 2,3,7,8 TCDD TO HUMANS FROM

FUGACITY LEVEL III DISTRIBUTION USING ROSENBLATT* CALCULATION

ONTARIO GUIDELINES

Acceptable level in fish
-2C ng/kg (consumed at rate
of 113 g/veek)

-results in exposure of
0.3 ng/day

-to reduce exposure to
conform with guidelines,
reduce emissions by
factor of 0.3/1.0 ,

Atinotpher«[

^ lb.

Temp<

ASSUMED

to
PARAMETERS ,

kvk

irature

25*C

Body

Weight

70 kg

\ \

Water Intake

2L/day

luman'

Food

Intake

0.63 kg/day

Food

Fraction of Diet

- fish

0.05

- meat

0.2

- 1 crop

0,1

*RosenblaCt, D.'H., Dacre, J.C., Cogley, O.K.,
in Environmental Risk Analysis for Chemicals,
Conway, R.J., ed. Van Nostrand Reinhold, N.Y.
1980.

- 191 -

CHEMICAL IDENTIFICATION AND BIOLOGICAL ASSAY
OF ENVIRONMENTAL MUTAGENS, PROMOTERS AND INHIBITORS

Morris Katz, K. R. Sharma and A. S. Raj

Centre for Research on Environmental Quality
York University, U700 Keele Street,
North York, Ontario M3J 2R3

- 192 -

Technology Transfer Conference No. 5,
Holiday Inn, November 27 & 28, 1984

CHEMICAL IDENTIFICATION AND BIOLOGICAL ASSAY OF

ENVIRONMENTAL MUTAGENS, PROMOTERS AND INHIBITORS

Morns Katz , K.R. Sharma and A.S. Raj
Centre for Research on Environmental Quality
York University, 4700 Keele St., Downsview, Ontario M3J 2R3

Abstract

Chemical Identification and Analysis of Polynuclear Aromatic
Hydrocarbons (PAH):- Earlier studies of the extraction and recovery of
PAH in water samples by adsorption on Cis Sep-Pak cartridges and analysis
by HPLC indicated that quantitative recovery was dependent upon pH of
the sample (Technology Transfer Conference No, 4). Consequently, the
effect of variation in pH of the water samples on the efficiency of
recovery of 16 PAH compounds, listed as high priority pollutants, was
determined over the range of pH 7, 5, 4, 3 and 2. The lower molecular
weight PAHs, consisting of 2 and 3 aromatic rings, can be recovered in
high yield at pH 7. Fluoranthene and pyrene can also be recovered in
yields of over 90 percent at pH 7. However, more complex PAHs of higher
molecular weight require adjustment of the water samples to pH 3 or 2
for optimum recovery. The influences of flow rate of water sample and
concentration limits of PAH on quantitative recovery by trace enrich-
ment through Sep-Pak cartridges were also investigated.

Biological Assay of Environmental Mutagens, Promoters and
Inhibitors:- The effects of four potential inhibitors, such as a-naphtho-
flavone. ascorbic acid, caffeic acid and ethoxyquin on pairwise combin-
ations of promutagens and direct-acting mutagens were determined by means
of the Ames Salmonella typhimurium test. The bone marrow micronucleus
assay on mice was employed to determine the inhibitory activity of
fumaric acid, thioacetamide, glutathione and 3-carotene on promutagens
and direct-acting mutagens in mice.

The effects of pairwise combinations of promutagens and direct-
acting mutagens in mice were studied by means of the bone marrow micro-
nucleus and abnormal spermhead assays. 7,12-Dimethylbenzanthracene (DMBA)
and benzo(a)pyrene (BaP), in combination, yielded a non-addtiive or
antagonistic response. Whereas additive responses were found with DMBA
plus cyclophosphamide (CP); BaP and CP; BaP and mitomycin C (MMC); CP and
MMC. In the sperm abnormality assay, pairwise combinations of DMBA plus
CP; and BaP+CP showed additive responses. However, non-additive responses
were produced by conti nations of DMBA+BaP; UMBA+t^lC; CP+MMC; and BaP+MMC.

- 193 -

Chemical Identification and Analysis of Polynuclear
Aromatic Hydrocarbons (PAH) in Water

1. Introduction

The use of reversed-phase chromatographic packing
to extract and concentrate trace levels of organics
from water has been described by several investigators
(1,2,3). Wolkoff and Creed (H) have reported an invest-
igation on the use of Sep-Pak C,g cartridges for the
collection and concentration of environmental samples .
A Sample Enrichment Purification cartridge (Sep-Pak) is
a product of Waters Associates and consists of a small
prepacked column containing 0. 35 g of C-. -PorasilA,
enclosed in a polyethylene cover which has been subjected
to radical compression to form a chromatographic bed.
In their limited study (Wolkoff and Creed, 4) only four
PAHs were determined and no mention was made of optimum
conditions for high recovery.

HPLC studies of PAH compounds have also been reported
by other investigators (5,6,7,8,9,10). However, data on
the efficiency of recovery and reproducibility of analytical
methods based on adsorption trapping and trace enrichment
of PAHs vary widely with different solid matrices. Katz
et al. ( 11 ) have reported earlier on the recovery of

- 194 -

PAH compounds (listed as high priority pollutants) from
water samples adjusted to pH 2, with efficiences in the
range of about 80 to over 90 percent, with a relative
standard deviation of less than 10 percent. Eichelberger
et al. (12) reported similar results of high recovery of
these high priority pollutant PAHs from water samples
adjusted to pH 2, although their results varied by a
relative standard deviation of 20 percent.

Losses of PAH may occur by adsorption on glass and
metal surfaces and constitute a problem mentioned by
several other investigators (9,13). It is apparent that
the use of Sep-Pak cartridges for trace enrichment of PAH
and careful manipulation to prevent adsorption losses in
tubing and pumping systems can provide meaningful data
by HPLC analysis.

- 195 -

Recently, re versed-phase liquid chromatograph packing
has been used to extract and concentrate trace levels of
organics from water. May et al. (14) used a C^ Q-bonded
phase packing for the extraction of hydrocarbons from
water. Traditional concentration and cleanup techniques
employ multiple solvent extractions followed by solvent
evaporation. These time consuming methods usually employ
relatively large amounts of solvents . There is , however,
an urgent need for rapid methods because in many cases a
large number of samples have to be analyzed. The technique
of trace enrichment using a suitable column affords a
simple means of simultaneous extraction, concentration
and purification of trace amounts of organics from water
samples in the same step.

Concentrations of polynuclear aromatic hydrocarbons
(PAHs ) in drinking water sources and supplies have been
found to range between a few nanograms to several hundred
nanograms per liter (15,16,17). Separation of isomeric
PAHs at these low concentrations has been difficult with
packed column GC analysis using both conventional and MS
detectors (18,19). All these methods utilize liquid-liquid
extraction, the traditional method of extracting and concent-
rating organic compounds from water which require several
time consuming steps, large volumes of volatile organic

- 196 -

solvf^nts and considerable analytical skill. An alter-
native method is adsorption trapping, in which the
organic compounds are extracted from a flowing sample
by strong adsorption onto a solid matrix. Various
adsorbents have been proposed as effective matrices
for removal of organ ics from water (20,21,22).

More recently, reverse-phase octadecylsilane (ODS )
chromatographic packing has also been used as a means
of trace enrichment of organics from various aqueous
samples including distilled water (23), drinking water
(1 ,2U ) , chlorinated water (25 ) , natural water ( 26 ) and
waste water (27).

Analytical methods based on adsorption trapping and
trace enrichment of PAH vary widely with different solid
matrices . It was considered advisable , therefore , to
establish the isocratic conditions and extent of recovery
of PAH in water samples by means of Sep-Pak cartridges ,
over a wide range of pH from neutral to increasingly
acidic conditions.

2 . Materials and Methods

Chromatography was conducted using Waters liquid chroma -
tograph Model 204 with dual Model 6000A solvent delivery
pumps , a Model U6K universal injector, a Model 600 solvent
programmer, Supelco LC-PAH reversed- phase analytic column
and Waters radial compression separation system containing

- 197 -

Z-module and a Radial-Pak PAH analytical cartridge column.
Data recording was done on a Spectra Physics Model 42 70
Computing Integrator. Detection was accomplished via
Waters Model 440 UV detector at 254 nm.
Preparation of Standards and Samples

The PAH standards used in this study were obtained from
Supelco containing 16 PAHs which comprise the U. S . Environ-
mental Protection Agency (EPA) list of priority pollutants
for water analysis. These compounds were obtained as
solutions in methanol/dichlorometh'ane 50/50 (V/V) and were
diluted further, 1 ml to either 10 ml or 100 ml, in
acetonitrile prior to use as standards . The stock solutions
of standard PAHs were stored in a refrigerator. The
synthetic samples of water were prepared by adding 1 ml of
diluted standard PAH solution to 50 ml of high purity water
from Millipores-Milli Q purification system.
Procedure

Each Waters Sep-Pak cartridge contains 0.35g of C-, „ porous

2
(70 pm diam) packing material (140 m of surface area per

cartridge). The Sep-Pak obviated hand-packing of non-
disposable guard columns and thus reduced any accidental
contamination and packing variations associated with this
procedure. More importantly, Sep-Pak cartridges allowed the
use of hand-held syringes instead of electrical pumps to
pass through water samples for extraction of organics, thus
improving portability for on-site sampling. These cartridges

- 198 -

are made specifically to fit a Luer-lock tipped syringe
for introduction of solution or solvents manually.

Prior to trace enrichment, the Sep-Pak cartridge was
always activated by passing 10 ml tetrahydrofuran , 10 ml
methyl alcohol followed by a rinse of 10 ml of water at
the (!esired pH. Then 500 ml of synthetic water sample
at a PAH concentration of 16 ng/ml at the required pH was
passed through the activated Sep-Pak at a flow rate of
55.6 ml per min. The Sep-Pak was dried by passing dry
nitrogen gas through it. The adsorbed PAHs were eluted
with 10 ml of glass-distilled dichloromethane into a
?5 ml pear-shaped flask. Dichloromethane was removed
completely from the extract with a stream of dry nitrogen
by a rotary evaporator. One ml of glass-distilled aceto-
nitrile was then added to the residue and 25 pi of this
sample was analysed by HPLC, using Waters Z-module consisting
of a Radial Pak PAH column and CH^ CN/H^O solvent mixtures.
at 7.0 ml /min .

Runs were carried out under both isocratic and gradient
conditions over the pH ranges of 2,3,U,5, and 7, in order
to determine the efficiency of PAH recovery. Quantitation
was accomplished by comparing the sample peak areas to
those of standards obtained under identical conditions.
Trace concentration of PAHs can be separated into subgroups
as fol lows , under isocratic conditions :

('i) Two and three membered ring systems separated by
60/UO and 50/50. V/V, CH^CN/H^O.

(b) four membered ring systems separated by 75/25
CH^CN/H^O.

- 199 -

(d) PAHs of higher molecular weight separated by
95/05 CHgCN/H^O or 100% CH^CN.

3. Results

The identity of the 16 PAH standards (high priority)
compounds and their HPLC retention times, in order of
elution, are listed in Table 1.

The data in Table 2 Indicate the efficiency of recovery
of 16 priority PAHs from pH adjusted water samples after
trace enrichment by C^ ^ Sep-Pak cartridges and subsequent
analysis by HPLC. It is noted that at pH7 the recovery
efficiency is high for the first 8 PAH compounds, from
naphthalene to pyrene. Thereafter the recovery for the
remaining 8 PAH decreases markedly. The recovery of
naphthalene decreases rapidly with increasing acidity of
the water samples. However, the remaining 15 PAH compounds
show improved recoveries as the pH of the samples is
decreased to pH3. At pH2 , high recoveries of PAH are
maintained, except for naphthalene, acenaphthylene and
acenaphthene .

Efficient recoveries of benzo (a) anthracene, chrysene,
benzo (b) f luoranthene , benzo (k) fluoranthene , benzo (a)
pyrene and PAH of higher molecular weight can be accomplished
only by acidification of the water samples to pH3 or 2 for
trace enrichment purposes . This phenomenon is due probably

- 200 -

to the electron density configurations of the high

molecular weight PAH compounds.

TABLE 1 : Retention Time of PAHs (Supelco Standard) and
Their Concentration

PAH Compounds Retention Concentration
(in order of elution) (min) (mg/ml)

1 Naphthalene Nph

2 Acenaphthylene Acelene

3 Acenaphthene Ace
'4 Fluorene Fl

5 Phenanthrene Phe

Fj Anthracene An

7 Fluoranthene Ft

8 Pyrene Py

9 Benz (a) anthracene B(a)an

10 Chrysene Chy

11 Benzo(b)f luoranthene B(b)Ft
1? BenzoCk ) fluoranthene B(k)Ft
13 Benzo(a)pyrene B(a)Py

1 N DibenzCa ,h) anthracene diB(a,h)An

1 'j Benzo(g,h , Dperylene B(g,h,i , )Per

IG IndenoCl,2 ,3-cd)pyrene In (1 , 2 , 3-cd)Py

4.49

0.10

6.06

0.20

8. 75

0.10

10.14

0.02

13.44

0.01

16. 04

0.01

18. 16

0.02

19. 12

0.01

23. 57

0.01

24.18

0.01

27.26

0.02

28. 70

0.01

29.60

0.01

32. 14

0.02

32.64

0.02

33.71

0.01

- 201 -

TABLF. ?: % Recovery of 16 priority PAHs by Sep-Pak C^
cartridges from water samples adjusted to
indicated pH levels.

Number

PAH

pH2

pH3

pH4

pH5

pH7

Nph

2.30

4.90

1.70

52.4

84.2

Acelene

23.5

92.2

85.0

79.1

94.3

Ace

69.9

100.0

88. 7

79.4

95.1

92.1

101.2

86.4

74.2

97. 7

Phe

94.9

104.0

100.0

85.2

97. 1

90.0

96.9

95.0

81.7

91.2

89.3

100.2

112.0

80.2

92.7

83.9

79.9

75.7

97.1

94. 3

B(a)An

87.1

85.4

75.1

53.0

60.1

Chy

87.8

86.0

64.1

39.1

44.6

B(b)Ft

86.5

88.4

60.9

38.7

40.4

B(k)Ft

85.8

82.4

51.6

21.0

45.6

B(a)Py

78.4

77.2

54.4

14. 3

41.2

diB(a,h)An

110.0

108.9

45.8

43.9

28.5

B(g,h,i)Per

86. 8

85.8

48.1

26.4

28.2

In(l,2,3-cd)Py

88.7

90.0

50.7

25.0

29. 3

- 202 -

H. Effect of Flow Rate on PAH Recovery

Several runs were made to investigate the effect of
increasing flow rate of the water samples during adsorption
of PAH by Sep-Pak cartridge on subsequent recovery of the
PAH, after elution with dichloromethane . The results
for fluoranthene , representative of a U-ring PAH system,
and for benzo (b) fluoranthene , a 5-ring PAH, are
presented in Table 3. The flow rate of the sample
solution through the Sep-Pak was controlled by means of
a Masterflex peristaltic pump. From the results present-
ed in Table 3 it is evident that there is no significant
variation in PAH recovery over the flow rate range of
2.0 to 2 8.6 ml/min.

Table 3: Effect of sample flow rate on recovery of
PAH from Sep-Pak

(concentration of PAH = 0.4 Pg/ml
pH of sample = 3.0)

n uoranthene Benzo (b) fluoranthene

Flow Rate % Recovery Flow Rate % Recovery
ml/min ml/min

, 7

28.

47.

. 1

57,

B3,

75.4 2...Q 81.0
76.1 9.7 81.1

78.5 28.6 82.4
72. 5

72.5
73.9

- 203 -

Lflc'ct of Drv-ing and Extraction of Loaded Sep-Pak
on PAH Recovery

The methodology developed for the trace concent-
ration and recovery of PAHs in water samples involves

(a) drying of the C^g Sep-Pak cartridge with dry
nitrogen gas after loading with the PAH sample, and

(b) extraction of the PAHs from the loaded Sep-Pak
with methylene chloride and evaporation of the CH^Cl^
extract. As both of these steps may involve potential
losses of PAHs , experiments were conducted to determine
the extent of PAH recovery in each case.

In separate experiments, two activated Sep-Pak
cartridges were loaded with an identical mixture of
PAHs in samples adjusted to pH 3, consisting of
naphthalene, pyrene , benzo(k) fluoranthene and benzo(a)
pyrene. In the case of one Sep-Pak cartridge, dry
nitrogen gas was passed through the cartridge to remove
moisture, prior to extraction of the PAHs and subsequent
recovery. In the other case , the loaded cartridge was
extracted first with 5 ml of methanol and then with
5 ml of tetrahydrofuran, without first drying the
cartridge with nitrogen gas after loading.

As indicated in Table 4, the use of nitrogen gas
for drying purposes has little effect on the recovery
of the high molecular weight PAHs, whereas the comp-
aratively volatile naphthalene suffers a pronounced loss

- 204 -

Table H : Effect of drying of loaded Sep-Pak on PAH recovery
(Samples adjusted to pH 3)

Compound

Concentration Percent recovery Percent recovery
pg/litre after drying solvent extraction
with N^

Naphthalene 200

Pyrene 2

BenzoCk ) f luoranthene 2

Benzo (a) pyrene 2

5.1
0. 1

12.5
10.2

14. 8
80.6
84.1
79.4

An examination was made of the possible loss of PAHs in
dichloromethane extracts of 16 U.S. EPA high priority PAH
pollutants during evaporation of the CH2CI2 extracts in a
50-inl Kuderna-Danish rotary evaporator. The PAH compounds
and their concentrations (pg/ml) in the CH^Cl^ extracts were
as follows:- naphthalene, 100; acenaphthylene , 200; acenaph-
thene, 100; fluorene, 20; phenanthrene , 10; anthracene, 10;
fluoranthene, 20; pyrene, 10; benz(a)anthracene , 10; chrysene, 10;
benzo(b)fluoranthene, 20; benzo(k)f luoranthene , 10;
benzo(a)pyrene, 10; dibenz(a,h)anthracene , 20; benzo(g,h,i , )-
perylene, 20; and indenoCl ,2 , 3 ,4-cd)pyrene , 10. Recoveries of
these 16 PAHs were determined by our HPLC method, as described
above, at pH3. With the exception of the comparatively
volatile naphthalene which showed a recovery of only about
21 percent, all of the other 15 PAHs were recovered in the
range of about 91 to 100 percent.

6. Recovery of Low Concentrations of PAH in Water Samples

Having optimized the conditions for the trace concentration
and recovery of PAHs by use of Sep-Pak cartridges, it was

- 2 OS -

considered advisable to extend the methodology to

lower concentrations of PAHs in water in order to

test the applicability of the procedure. The results

of a study of the recovery of PAHs at concentrations

of 1 to 2 yg/litre, for the majority of the compounds,

are presented in Table 5. Most of the compounds were

recovered to the extent of about 80 percent or higher.

However, some compounds were recovered in yields of

less than 70 percent. Such losses were investigated

further to determine whether they could be due to

significant adsorption on the walls of tubing and

pumping system.

Table 5: Recovery of PAHs in low concentration in
water using Sep-Pak cartridges and HPLC
analysis (solution pH adjusted to 3.0)

Compound

Concentration
yg/litre

Percent Recovery

Acenaphthylene

Acenaphthene

Fluorene

Phenanthrene

Anthracene

Fluoranthene

Pyrene

Benz(a)anthracene

Chrysene

Benzo(b) fluoranthene

BenzoCk) fluoranthene

Benzo (k ) fluoranthene

0.01

Benzo(a)pyrene

BenzoCa)pyrene

0.02

DibenzC a, h) anthracene

Benzo(g,h,i)perylene

IndenoCl , 2 , 3-cd) pyrene

74. 3
69.0
77.1
83.0
88.0
84. U
70.4
77. 8
77.6
68.1
81.6
83.4
96.5
70.4
66. 8

65.6
54.7

- 206 -

Therefore, several experiments were carried out
with synthetic water samples containing fluoranthene
(80 yg/L) and benzo(a)pyrene (20 ug/L) using a small
steel pre-analysis extraction column (4.6 mm x 5.0 cm)
containing the same packing material as Sep-Pak for
trace concentration of the PAH. When this small
column was loaded with 10 ml of the water sample at
pH 3.0, flushed with eluting solvent and analyzed by
our usual HPLC procedure, comparatively lower recoveries
were obtained than with use of a Sep-Pak cartridge.
However, rinsing of the feed tubing system and wetted
parts of the pump with 100% acetonitrile increased the
recovery to about 100 percent for fluoranthene and
98.4 percent for benzo(a)pyrene.

7, Analysis of FAHs in Tap Water

2.0 litres of tap water, adjusted to pH 3.0 with
O.IN HCl, were passed through an activated Sep-Pak
cartridge at a flow rate of 13 ml/min for trace
concentration of PAHs . The Sep-Pak was dried by
passage of dry nitrogen gas and then eluted with
5 ml of dichloromethane. After evaporation of the
solvent in a Kuderna-Danish rotary evaporator to 1 ml
and then to dryness with a stream of dry nitrogen gas,
the residue was dissolved in 0.5 ml of pure acetonitrile

- 207 -

■1

An aliquot of 25 vl of sample was analyzed by
HPLC as described above. Quantitation was accomp-
lished by comparing the sample peak areas with those
of standards under identical conditions. The results
are presented in Table 6 as the averages of 3 analyses

Table 6 : Analysis of PAH concentrations in tap water

Compound identified
by HPLC

Fluoranthene
Benzo(b)fluoranthene
Benzo(k) fluoranthene
Benzo(a)pyrene
Indeno(l,2 ,3-cd)pyrene

Concentration

ng/litre
(average of 3 analyses)

4.0
2.0
1.1
4.7
1.1

- 208 -

Re f erf^nces

1. Ogan, K,. Katz, E. and Slavin , W. J. Chromatog. Sci
16:517 (1978).

2. Cartwell, F.F., Anal.Chem . 48:1854-1859 (1976)

3:. Eisenbeiss, V.F., Hein, H. , Joster, R. and

Naundorf, G., Chromatogr. Newslett. 6:8-12 (1978)

4. Wolkoff, A.W. and Creed, C. , J. Liq. Chromatogr.
4(8) :1459-1472 (1981)

5. Crosby, N.T., Hunt, D.C., Philp, L.A., Patel, I.,
Analyst (London) 106 : (1259 ): 135-45 (1981)

6. Roumeliotis, P. and Unger, K.K., Anal. Chem. Symp,
Ser . 3:229-45 (1980)

7. Joe, Frank L. , Roseboro, Emina L. , Fagio, Thomas,
J. Assoc. Off. Anal. Chem. 64 ( 3) : 641-646 (1981)

8. Durand, J. P. and Petroff, N. , J. Chromatogr.
190:85-95 (1980)

;9„ Ogan, Kenneth and Katz, Elena, J. Chromatogr .
188:115-127 (1980)

10. Wise, Stephen A. and May, Willie E., Anal. Chem .
55(9) :1479-1485 (1983)

11. Katz, M. , Raj, A.S. and Sharma , K.R., Proceedings
Technology Transfer Conference No. 4, pp. 216-240,
Ontario Ministry of Environment, Toronto,
November 29-30, 1983.

12. Eichelberger, J.W., Kerns, E.H., Olynyk, P. and
Budde, W.L., Anal. Chem. 55:1471 (1983)

13. Karasek, F.W., Clement, R.E. and Sweetman, J. A.,
Anal. Chem . 53(9) :1050A (1981)

14. May, W.E., Chesler, S.N., Cram, S.P., Gump, B.H.,
Hertz, H.S., Enagonio, D.P. and Dyszel, S.M.,

J. Chromatogr. Sci . 16:517 (1978)

15.. Lewis, W.M., Water Treat. Exam. 24:243 (1975)

- 209 -

16. Saxena, J., Besu, D.K. and Kozuchowski, J., Health
Effects Research Laboratory TR-77-563, No. 24,
Cincinnati , Ohio , 1977

17. Sorrel, R.K. , Reding, R. and Braes, H.J., 177th
National Meeting of American Chemical Society,

19 Div. of Enviro. Chem. , Honolulu, Hawaii, Paper
No. 126, 1979

IS* Giger, W. and Shaffer, C. Anal. Chem. 50:243 (1978)

19. Lee, M.L., Vassiloros, D.L., Pipkin, W.S. and
Sorsen, W.L. , Trace Organic Analysis, National
Bureau of Standards, Special Publications 519,
Washington D.C., 1979, p. 731

20. Navratill, J., Sievens, R. and Walton, H. , Anal. Chem .
49:2260 (1970)

21. Saxena, J., Kozucholowski , J. and Basu, D. Environ.
Sci. Technol. 11:682 (1977)

22. Thurston, A.D. , J. Chromatogr. Sci . 16:254 (1978)

23. Elsenbeiss, F. , Hein, H. , Jacoter, R. and Naundorf, C. ,
Chromatog. News 6:8 (1978)

24. Waters Associate Technical Bulletin H63, November 1976

25. Oyler, A.R., Bodenner, D.L. , Welch, K.J., Liukkonen, R.J,,
Carlson, R.M., Kopperman, H.L. and Caple, R. Anal.

Chem . 50:837 (1978)

26. Kummert, R. , Molnar-Kubica, E. and Giger, W. Anal. Chem .
50:163 (1978)

27. Waters Associates Technical Bulletin H91, October 1977

- 210 -

BIOLOGICAL ASSAY STUDIES OF ENVIRONMENTAL MUTAGENS,
PROMOTERS AND INHIBITORS

1. Introduction

The studies of mutagens, potential inhibitors and promoters
have been conducted by using a three-tier system consisting of
the in vitro Ames Salmonella typhimurium assay, (Ames et al.,
1975) , the in vivo bone marrow micronucleus assay (Heddle and
Salamone, 1981), and the abnormal spermhead assay (Wyrobeck and
Bruce, 1975) with mice. This tier system has been employed to
study the influence of pairwise combinations of promutagens and
direct-acting mutagens in order to determine additive, antagonistic
and/or synergistic effects. Potential inhibitors such as fvimaric
acid, thioacetamide, glutathione and 3-carotene have been examined
to determine their influence in pretreatment of mice subsequently
subjected to the action of single mutagens and assessed by the
bone marrow micronucleus assay. The effects of fumaric acid and
thioacetamide on mutagenic activity were also assessed on mice by
the abnormal spermhead assay.

The Ames Salmonella typhimurium assay was also employed to
determine the inhibitory effect of a-naphthof lavone , ascorbic acid,
caffeic acid and ethoxyquin on pairwise combinations of promuta-
gens and direct-acting mutagens.

- 211 -

2. The Effects of Pairvise Combinations of Mutagens on Mice

The human population is exposed to an environment which con-
tains a mixture of pollutants, especially in the urban air where
contaminants, including polynuclear aromatic hydrocarbons (PAH),
are generated by incineration of refuse, combustion of fossil
fuels, exhaust from motor vehicles and from many other sources.
PAH are relatively inert per se until they are metabolized by
mammalian monooxygenase enzymes to biologically active products
in order to become carcinogenic or mutagenic. These active compounds,
or ultimate carcinogens, are diol epoxides, in which the epoxide
moiety forms part of the bay region of the hydrocarbons (Conney
et al., 1978; Jerina et al., 1980). These PAH are not present
individually but as coit5)lex mixtures.

The effect of such mixtures on the health of exposed human
beings is an important aspect that requires invetigation. Would
the different PAH interact and what sort of action would they
have on human population? Would there be synergistic effects
or less than additive effects? In order to learn these basic
facts, experiments have been conducted using different combinations
of promutagens as well as a direct-acting mutagen. A modified in
vivo bone-marrow micronucleus assay, as described by Heddle and
Salamone (1981) and Salamone et al. (1980), has been used in
order to detect the chromosomal breaks in the somatic cells .

- 212 -

2.1 Methods of Pairwise Combinations on Somatic Cells

Promutagens 7, 12-dimethylbenz(a) anthracene (DMBA) , benzo(a)
pyrene (BaP) and cyclophosphamide (CP) and a direct-acting mutagen,
mitomycin C (MMC) were employed.

Virgin B6C3F1 female mice were 8-9 weeks of age when used
for the in vivo bone marrow micronucleus assay experiments. The
mice ranged from 20-22g in weight and they were fed with Purina
Laboratory Rodent Chow end watered ad libitum.

Mice were given single treatments (intraperitoneal injections)
of the appropriate mutagen. The mice that received two mutagens
in combination, each mutagen was administered individually within
an interval of 10-15 minutes. The individual treatments included
DMBA (30mg/kg in DMSO) , BaP (186mg/kg in DMSO) , CP (45mg/kg in
physiological saline), and MMC ( Img/kg in physiological saline).
The pairwise combinations included DMBA+BaP, DMBA+CP, DMBA+MMC,
BaP+CP, BaP+MMC, and CP+MMC.

4-5 mice were included in each treatment per each sampling
time. Bone marrow samples were collected at various intervals
(24, 48, 72 hr and in some cases, 96 h) and 500 polychromatic
erythrocytes (PCE) were scored at each sampling interval per
animal and the results expressed as the average number of micro-
nucleated PCE.

- 213 -

2.2 Results of Pairwise Combinations on Bone Marrow

Untreated mice showed an average spontaneous frequency of
micronucleated PCE ranging from 0.4 to 0.6 per 500 PCE . Other
solvent controls showed an average 0.4 (NaCl) to 1.6 (DMSO)
micronucleated PCE per 500 PCE. The results from individual treat-
ments as well as pairwise combinations are presented in Figs. 1-6
and some salient points are summarized in Table 1. It is observed
that when two promutagens of similar mode of activation (DMBA and
BaP) are combined, the response is non-additive. When two pro-
mutagens of different modes of activation (DMBA and CP; BaP and CP)
are employed in a pairwise treatment, the response is additive.
Similar "additive" responses were also observed when a promutagen,
either BaP or CP, and direct-acting mutagen MMC were involved in
combination. However, with promutagen DMBA and direct-acting
mutagen MMC, the response was variable, as the effect was inter-
mediate in one series of tests and additive in earlier tests.

The results obtained from these experiments show some of the
influence of enzyme activation systems on mutagenicity. Since
PAH are metabolized by a microsomal mixed-function oxidase enzyme
called aryl hydrocarbon hydroxylase (AHH) , the resultant activation
process may lead to the formation of an intermediate epoxide
species which is a much more active carcinogen than the parent
hydrocarbon. The epoxide can be converted into a glycol or diol

- 214 -

by a second microsomal enzyme, epoxide hydrase, which may lead

to detoxification and rapid removal of the reactive epoxide before

interaction with a target site.

When promutagens DMBA and BaP, which are activated by similar
metabolic pathways, were tested in combination, the mixture
showed less than an additive response. Additive response was
noticed where the promutagens in a pairwise combination were
activated by different pathways, as is seen in the combination of
either DMBA and CP, or BaP and CP. Our results confirm the earlier
results of Salamone et al. (1982). However, they reported the
results from only one sampling point. Additive response was
also observed when the direct-acting mutagen MMC was involved in
the combination with either BaP or CP.

Since the reactions with the promutagens D^BA, BaP, and CP
require metabolic activation to show mutagenic activity, a respon-
sive strain of mice B6C3F1 was used in the present work because
the monooxygenase enzyme system in this strain can be induced by
PAH (Nebert et al., 1972; Green, 1973). MMC is a direct-acting
agent and it is capable of acting as a conventional alkylating
agent (Iyer and Szybalski, 1964) .

Possible mechanisms could be proposed regarding the mode of
action of the mutagens. If the mutagens have similar modes of
action, they compete for the same enzymes to become activated and
they may not obtain enough enzymes from the system. So, instead of

- 215 -

an additive effect, there is fortunately, a lesser effect that
is comparable to either promutagen acting singly, e.g., DMBA+BaP.
If the mode of action involves different enzyme systems or meta-
bolic pathways (e.g., BaP and CP) , the mutagens do not have to
compete for activation enzymes. Therefore, one can expect an
additive reaction,

Table 1: Mutagenic Activity of Pairwise Combinations of Model
Agents on In Vivo Bone Marrow Micronucleus Assay

Pairwise
Combination

Types of
Mutagen

Mode of
Activation

Response

DMBA+BaP

DMBA+CP
DMBA+MMC

BaP+CP
BaP+MMC

CP+MMC

Promutagens

Promutagen &
direct-acting

Promutagens

Promutagen &
direct-acting

Similar modes

Different modes
Different modes

Promutagen &
direct-acting

DMBA - 7 , 12-dimethylbenz (a) anthracene
BaP - Benzo(a) pyrene
CP - Cyclophosphamide
MMC - Mitomycin C

Non-additive (less
than additive)

Additive

Non-additive ( inter ■
mediate effect)

Additive

Different modes Additive

- 216 -

• • DMBA

A ^BaP

■ ■ DMBA *BaP

o
o

O
5

t'.

~48

I —

Figure 1

HOURS

Frequency of mlcronucleated PCE/500 PCE in mice
bone marrow as a function of time after treating
with DMBA, BaP and combination of DMBA+BaP.

- 217 -

78-

16-

14-

12-

6-1

O O DMBA

A---A DMBA -{-CP

HOURS

Figure 2

Frequency of micronucleated PCE/500 PCE in mice bone
marrow as a function of time after treating with
DMBA, CP and combination of DMBA+CP.

- 218 -

78-

16-

14-

Uj 12-

O
O

uj
O

4-1

O owe /I

A /WA//C

DMBA +• M/WC

----4

— T"

~ni —
96

HOURS

Figure 3:

Frequency of micronucleaCed PCE/500 PCE In mice bone
marrow as a function of time after treating with
DMBA, MMC and combination of DMBA+MMC.

- 219 -

O
O

22-

20-1

16-t

16-

14H

12'

10-

O O BaP

A -▲ C P

4 ♦ BaP-h CP

~i —

— I —

HOURS

Figure 4:

Frequency of mlcronucleated PCE/500 PCE In mice bone
marrow as a function of time after treating with
BaP, CP and combination of BaP+CP.

- 220 -

-O Bap

18-

16-

14-

12-

Uj 10-

1
o

^ ft-l

in On

A -A mmC

BaP-hMMC

Figure 5

Frequency of micronucleated PCE/500 PCE in mice bone
marrow as a function of time after treating with
Bap, MMC and combination of BaP-l-MMC

- 221 -

Uj
O

HOURS

Figure 6: Frequency of micronucleated PCE/500 PCE In mice bone
marrow as a function of time after treating with
CP, MMC and combination of CP+MMC.

- 222 -

2.3 Methods of Pairwise Combinations on Germ Cells

The effect of mixtures of PAH on exposed human beings is
important from the viewpoint of genetic risk. Those PAH which
induce transmissible sperm damage in males might lead to terato-
genic events. Therefore, we have extended our experiments to germ
cells of mice in studying the effect of mixtures of mutagens
using the sperm abnormality assay of Wyrobek and Bruce (1975) .

B6C3F1 male mice were 11 to 14 weeks of age when used for the
sperm abnormality assay. The mice ranged from 24-27g in weight
and were maintained under good animal room conditions before and
during experimental periods of 35 days.

Promutagens, DMBA (25mg/kg), BaP (20mg/kg), and CP (45mg/kg)
and direct-acting MMC ( . 8mg/kg) were injected i.p. individually
and in pairwise combinations. Dimethyl sulfoxide (DMSO) was the
solvent for DMBA and BaP, whereas physiological saline was used
as solvent for CP and MMC. In experiments with single mutagens,
6-8 mice were used in each case, whereas in the pairwise combina-
tions ten mice were treated.

At least 3 injections of MMC at 0.8mg/kg were necessary in
order to obtain meaningful data. Therefore, MMC was injected once
every 24 h for 3 days consecutively. With the other 3 mutagens,
single injections were sufficient. In pairwise combination experi-
ments, the total dose of MMC was 2 . 4mg/kg (i.e., 3x0.8mg/kg).
Sperm cells were collected frcxn Cauda epididymides on the 35th

- 223 -

day after the last injection and sperm smears were prepared
according to Wyrobek and Bruce (1978) , with slight modifications
(Raj and Katz, 1984) . Abnormal sperm cells were counted (Wyrobek
and Bruce, 1975) per 500-1000 sperm cells at 400x magnification.

2.4 Results of Pairwise Combinations on Germ Cells

Experimental results are plotted in the form of histograms
(Figs. 7-10) and a summary is presented in Table 2.

Maximum numbers of abnormal sperms were observed in treatments
with direct-acting mutagen MMC. Non-additive responses were ob-
tained in combination treatments of DMBA+MMC, CP+MMC, DMBA+BaP
and BaP+MMC. Additive response was observed in combinations of
promutagens DMBA+CP, and BaP+CP. The in vivo sperm abnormality
assay results concurred with in vivo bone marrow nucleus assay
results in four combination treatments, viz., DMBA+BaP, DMBA+CP,
DMBA+MMC and BaP+CP . The results differed in only two cases,
BaP+MMC and CP+MMC, where the effects in germ cells were a non-
additive response compared with an additive response in bone
marrow. In germ cells this constitutes a very favourable factor.

Wyrobek et al., (1983) reported that the sperm abnormality
assay could be useful in assessing hazards caused by chemicals.
Agents that induce increases in spermhead damage in mice are highly
correlated with known germ cell mutational activity. Chemicals
that are positive in the sperm abnormality assay should be con-
sidered for study in human population.

- 224 -

Table 2: Effect of Pairvise Combinations of Mutagens

- In Vivo Sperm Abnormality Assay with Mice -

Pairwise Combinations Effect

DMBA+CP Additive response

BaP+CP Additive response

DMBA+MMC Non-additive

CP+MMC Non-additive

DMBA+BaP Non-additive

BaP+MMC Non-additive

- 225 -

DM8A

DMBAi-BaP

Ba P

Figure 7: Sperm abnormality assay results

showing the effect of DMBA, DMBA+
BaP, and BaP.

- 226 -

60-

53-4

50-

a
m

o
o

E
a

40-

422

30-

20-

10-

31-5

14-6

13-5

27.2

47-8 47.2

334

40 3

DMBA CP DMBA+CP MMC DMBA+MMC

Figure 8: Sperm abnormality assay results showing
the effect of DMBA, CP , DMBA+CP. MMC and
DMBA+MMC .

- 227 -

o
c
-o

c
ctl

t53-4
47 8

40-

42. 2

30-

20-

10-

-1-49
-—,45

141

-r29

11-3

J-6:

90
6-5

2-5

14J-

21-5

MMC

+
C P

C P

BaP

CP-\-BaP

Figure 9: Sperm abnormality assay results showing the
effect of MMC, MMC+CP. CP, BaP and CP+BaP.

- 228 -

1U(J-

l99 L'

80 H

.5
lu

'^ 60-

BE
Q

40-

20-

-r 83 2

76 1

T 22 3
18 5

147

87 7

762

Ba P

BaP+ MMC

MMC

Figure 10: Sperm abnormality assay results

showing the effect of BaP, BaP+MMC,
and MMC.

- 229 -

2.5 Dose Response Studies of Mutagens in Ames Salmonella Assay
The dose response studies were conducted with each mutagen

in the presence of S-9 metabolic activation mixtures, 0.25ml
of 10%. BaP dose response was tested with 0.125, 0.25, 0.5 and
l.Oyg/plate; DMBA with 0.1, 1.0, 10.0 and 100 . Oyg/plate, using
tester-strain TA-98. The solvent was 0.1ml DMSO. Dose responses
for MMC were tested with 50 and lOOng/plate and for DMBA, 0,5, 5.0,
50.0 and 500 .Oyg/plate, using tester strain TA-102. CP dose
response was determined at concentrations of 0.1, 1.0, 10.0, 100.0
and lOOO.Opg/plate and MMC with 50.0 and 100. Ong/plate, using
tester strain TA-102. Strain TA-102 was constructed by Maron
and Ames (1983) primarily for detecting mutagens that require an
intact excision repair system. This strain detects efficiently a
variety of mutagens, and cross-linking agents, such as psorolens
and mitomycin C.

2.6 Results of Pairwise Combinations on Salmonella

The in vitro tests of pairwise combinations included the
mutagen pairs and effects listed in Table 3, in treatments with
dose levels listed above. All possible pairwise combinations
of doses for each mutagen pair have been included in the assay
program. The results for the combination of BaP and DMBA in terms
of the number of revertants serve to illustrate the non-additive
effect that was observed for this pair.

- 230 -

Table 3; Effect of Pairwise Combinations of Mutagens
Ames Salmonella Typhimurium Assay ,

Pairwise Combination Strain + S-9 Effect
of Mutagens TA-

BaP+DMBA 98 Non-additive,

antagonistic

DMBA+MMC 10 2 Additive

CP+MMC 102 Additive

BaP+CP 98 Additive

For example, in the case of BaP with LOyg/plate dose,
the number of revertants was 107, and for DMBA with lOOpg dose
per plate, the number of revertants was 57. Consequently, the
expected number of revertants would be about 160 for the combination
if the combined effect was additive. However, the actual number
of revertants, 68±2, indicated that there is an antagonistic
action or else the activity of the monooxygenase enzyme system
in S-9 is induced preferentially by DMBA and BaP is not activated.

The non-additive effect when two promutagens are involved is
possibly due to their similar mode of action. When the mutagens
have similar modes of action, they compete for the same enzymes
in order to become activated and they may not obtain enough enzymes
from the system. So, instead of an additive effect, there is
fortunately, a lesser effect that is comparable to either promutagen
acting singly.

- 231 -

The experimental results from pairwise combination of direct-
acting mutagen mitomycin C (MMC) and promutagen DMBA yielded
an additive number of revertants, as illustrated in Table 3 and
also by the following example. Treatment of TA-102 with lOOng
MMC per plate yielded 1100 revertants. DMBA treatment at a dose
of 5.0ug/plate yielded 246 revertants. The pairwise combination
produced an additive total of 1396 revertants. MMC is a direct-
acting alkylating mutagen (Iyer and Szybalski, 1964) .

3. Effect In Vitro of Potential Inhibitors Against Pairwise
Combinations of Mutagens

Alpha-naphthof lavone (a-NF or 7, 8-benzof lavone) , ascorbic

acid, caffeic acid, and ethoxyquin were the inhibitors tested

against pairwise combinations of mutagens in the Ames Salmonella

assay.

3.1 Alpha-naphthof lavone (a-NF)

Studies with flavones such as a- and g-naphthof lavones have
shown that animals can be protected from some of the deleterious
health hazards caused by polynuclear aromatic hydrocarbons (PAH) .
It was observed that a-NF acted as a potent inducer of increased
mixed function oxidase activity, resulting ultimately in inhibiton
of epidermal neoplasia caused by DMBA (Gelboin et al., 1970; Slaga
and Bracken, 1977; Wattenberg, 1979) . The mutagenic and clasto-
genic activity of DMBA was found to be inhibited by a-NF (Raj
and Katz, 1983, 1984) .

- 232 -

From the above reports it was evident that a-NF acts as an
inhibitor against PAH-induced clastogenicity . However, it is not
known how ot-NF acts against pairwise combinations of mutagens.
So we tested a-NF against pairwise combinations of BaP+DMBA,
DMBA+MMC, BaP+CP, and CP+MMC in the iji vitro Ames Salmonella assay.
The results are presented in Table 4. From the data in this table
one can notice that a-NF showed an inhibitory effect against
pairwise combinations of BaP+DMBA and DMBA+MMC. The effect of
a-NF against BaP+DMBA is similar to the pairwise combination by
itself. No inhibitory effect was noticed either against BaP+CP
or CP+MMC.

3.2 Ascorbic Acid (AA)

Effect of ascorbic acid (AA) against the pairwise combination
of mutagens was studied also in Ames Salmonella assay. The results
are presented in Table 4 .

Inhibitory effects were observed against the pairwise combina-
tions of BaP+DMBA, DMBA+MMC and CP+MMC. However, no significant
effect was observed against the pairwise combination of BaP+CP
at low concentrations of AA but some reduction in revertants
was evident at higher AA concentrations.

- 233 -

Table 4: Effect of Inhibitors on Pairwise Combinations of Mutagens
Ames Salmonella Assay

Inhibitor

Pairwise
Combination

+S-9

Strain

TA-

Effect

a-Naphthof lavone

BaP+DMBA

Inhibitory

DMBA+MMC

102

Inhibitory

BaP+CP

No significant effect

CP+MMC

102

No inhibition

Ascorbic Acid

BaP+DMBA

Inhibitory (similar to
combination of BaP+DMBA)

BaP+CP

Some inhibition at higher
doses

DMBA+MMC

102

Inhibitory

CP+MMC

102

Moderate inhibition

Caffeic acid

BaP+DMBA

Inhibitory (no significant
difference from BaP+DMBA

BaP+CP

Inhibitory in comparison
with combination

DMBA+MMC

102

Inhibitory in cou^arison
with combination

CP+MMC

102

Inhibitory

Ethoxyquin

BaP+DMBA

Inhibitory (similar to
pairwise alone}

BaP+CP

Inhibitory

DMBA+MMC

102

Inhibitory

CP+MMC

102

Inhibitory

- 234 -

3.3 Caffeic Acid (CA)

Caffeic acid or 3 , 4-dihydroxycinnaraic acid, which occurs
in many fruits and vegetables (Mozel and Hermann, 1974; Schmidtlein
and Hermann, 1975a, b; Stohr and Hermann, 1975; Stohr et al., 1975;
Pomenta and Burns, 1971) was found to inhibit the BaP-induced
neoplasia of the forestomach in mice (Wattenberg et al. , 1980)
as well as mutagenesis in vitro (Wood et al.) .

Recently, Raj et al. (1983) reported the inhibitory effect
of caffeic acid against DMBA-induced clastogenic action in mice.
Therefore, it was considered desirable to determine whether caffeic
acid inhibits in similar fashion a pairwise combination of mutagens.
Therefore, we tested the effect of caffeic acid against pairwise
combinations of BaP+DMBA, BaP+CP, DMBA+MMC and CP+MMC, in the Ames
Salmonella assay. The results from these experiments are also
tabulated in Table 4. Caffeic acid showed an inhibitory effect
in all the above-mentioned pairwise combinations.

3.4 Ethoxyquin (EQ )

Ethoxyquin, a non-phenolic antioxidant, was tested in Ames
Salmonella assay against the pairwise combinations of mutagens
BaP+DMBA, BaP+CP, DMBA+MMC and CP+MMC. The results are presented
in Table 4. Inhibitory effect was observed against all the mutagen
combinations. However, in the case of BaP+DMBA, the inhibition is
rather similar to BaP+DMBA as a pairwise combination alone in the
absence of EQ.

- 235 -

4. Results from Potential Inhibitors Assessed by In Vivo
Bone Marrow Micronucleus Assay

4.1 Fumaric Acid

Fumaric acid (C.H.O.) occurs in plants such as Caps el la
bur sa-pas tori s which can be used as green salad, and many other
plants including edible mushrooms. Kuroda and Takagi (1968)
and Kuroda et al. (1974) have indicated that extract of capsella
bur sa-pas tori s has various kinds of pharmacological activity
including antiulcerative and anticarcinogenic properties. Fumaric
acid was found to be a protective agent against 5-nitrofuran
(NFN) -induced forestomach and lung carcinogenesis in mice (Kuroda
et al., 1982). Fumaric acid was also found to be responsible for
inhibiting the growth of subcutaneous ly transplanted Erlich tumors
in mice (Kuroda et al., 1976) or gastric ulcers in rats (Kuroda
and Akao, 1977) . Furthermore, fuiaaric acid reduced the lethal
and hematological toxicity of mytomycin C (MMC) (Kuroda and Akao,
19 80) .

Since a protective effect of fumaric acid was observed against
the nitrogen-containing carcinogen, NFN, we wished to find whether
similar inhibitory effect of fumaric acid would be observed against
certain other nitrogen-containing mutagens, MNNG and MMC. In
addition, DMBA, a promutagen was chosen as another test mutagen
because DMBA is a potent clastogen in the in vivo bone marrow
micronucleus assay .

- 236 -

Experimental mice (B6C3F1 female) were fed with powdered
food containing 1% fumaric acid for one week prior to injecting
i.p. either with MNNG (50mg/kg in DMSO) , MMC (Img/kg in distilled
water) and DMBA (3 0mg/kg in DMSO) . Bone jTiarrow samples were
collected at 24, 48 and 72 hours. From each group of 5 to 7
mice per group, 1000 PCE were scored per animal and the results
are expressed as the average number of MNPCE/1000 PCE. Standard
errors of the mean (SEM) have been calculated between animal
samples .

The results from the above experiment are presented in Table
5. In the mice that were prefed with fumaric acid and that
received DMBA, the number of MNPCE increased both at 24 and 48 h
when compared with only DMBA treatment. MNNG did not show appre-
ciable change in the number of MNPCE in the presence of fumaric
acid. The other nitrogen-containing direct-acting mutagen, MMC,
showed a maximum nuiTiber of 23.4 MNPCE on average per 1000 PCE
at 24 h sample. However, at the same saitpling time, a reduction
in the number of MNPCE was observed (15.2 MNPCE/1000 PCE) in
the presence of fumaric acid indicating a substantial inhibitory
effect (see Fig. 11) .

DMBA requires metabolic activation to show mutagenic activity
This requirement was fulfilled by using a responsive strain
of mice, B6C3F1, since the monooxygenase enzyme system can be
induced by PAH in this strain {Nebert et al., 1972; Green, 1973).

- 237 -

If fumaric acid acts as an inhibitor of the monoxoygenase
enzyme, arylhydrocarbon hydroxylase (AHH) , that should reduce
the mutagenesis when given with DMBA . That was not the case.

4.2 Glutathione

Glutathione is a widely-distributed sulphur-containing tri-
peptide that consists of glutamic acid, cystein and glycine in
that order. The functional group in the molecule is the third
group and it is customary to represent reduced glutathione by
the abbreviation GSH.

Reduced glutathione is oxidized to the disulphide by mild
oxidizing agents, molecular oxygen, and by cytochrome C. It
is oxidized enzymatically by dehydroascorbate in the presence of
glutathione dehydrogenase; and enzymatically it can be reduced
by NADP or NADPH in the presence of glutathione reductase.
Since glutathione undergoes enzymatic oxidation and reduction,
it can act as a biological hydrogen carrier (Hopkins, 1921) .

Moir (1980) reported that reduced glutathione administered
to rats bearing aflatoxin B, -induced liver tumors caused re-
gression of tumor growth and resulted in survival of animals.

Mitchel and his colleagues (1973) reported the protective
role of glutathione in acetaminophen-induced hepatic necrosis
in mice. They observed that pretreatment of mice with diethyl

- 238 -

maleate, which depletes hepatic glutathione, potentiated aceta-
minophen-induced hepatic necrosis, whereas pretreatment with
cy stein, a glutathione precursor, prevented hepatic damage.

Hinson et al. (1981), in a mini review of acetaminophen-
induced hepatotoxicity, described how glutathione detoxified
the reactive metabolite of acetaminophen which was formed by a
cytochrome P-450 .

From the above reports it is clearly seen that glutathione,
especially in its reduced form, is a detoxifying agent, at least
against hepatotoxic agents. When an animal is treated with
known mutagens or carcinogens, it is believed that these chemicals
pass through the liver before reaching the target s ite ,

Experiments were conducted with mice by pretreating them
with glutathione (lOOmg/kg in distilled water) for 24 h prior to
treating either with DMB C (30mg/kg in DMSO) , or BaP (150mg/kg in
DMSO) . Bone marrow samples were obtained at 24 , 48 and 72 h .
The results are summarized in Table 5 and presented graphically
in Fig. 12.

4.3 3-Carotene

8-Carotene is the most important of the provitamins A. It
is widely distributed in the plant and animal kingdom. In plants
it occurs almost always together with chlorophyll.

- 239 -

There are several reports indicating that dietary vitamin A
(B-carotene) has a bearing on the reduction in the incidences of
lung cancer (Bjelke, 1975; Hyrayama, 1979; MacLennan et al., 1977;
Mettlin et al., 1979; Gregor et al., 1980; Shekelle et al., 1981).
Rettura et al . C1983) reported that diet supplemented with 3-
carotene (90mg/kg diet) prevented DMBA-induced tumors.

Experiments were conducted using g-carotene against DMBA-,
BaP-, CP- and MMC-induced chromosomal aberrations in mice, assessed
by the in vivo bone marrow micronucleus assay.

B6C3F1 female mice, 8 weeks old, were fed for one week on
powdered food containing B-carotene (lOOmg/kg) . Experimental
mice were injected i.p. with DMBA (30mg/kg in DMSO) , BaP (150
mg/kg in DMSO), CP (45 mg/kg in NaCl) ; or MMC (Img/kg in NaCl) .
Bone marrow samples were collected at various intervals after
the mutagen treatment. The experiments were repeated twice and
the results are presented in Table 5.

The results indicated an inhibitory effect on both DMBA-
and BaP-induced bone marrow MNPCE, In the case of mice treated
with either CP of MMC the results were not reproducible.

- 240 -

26-

24-

22-

20-

18-

16-

O. 14H

o
o
o

111

12-

10-

Q O FUMARIC ACID

MMC

MMC ♦FUMARIC ACID

-~l —

Ho u rs

Figure 11; Frequency of micronucleated polychromatic erythro-
cytes in bone marrow of mice as a function of time
after treating with either Mitomycin C, or Fumaric
Acid, or MMC pretreated with Fumajric Acid.

- 241 -

14-

^10-

o ..
o

Uj 8-
O

• DMBA

^ DMBA +G/ utathione

m Gl utathione

24 48

Sampling Time (h)

Figure 12

Frequency of mlcronucleated polychromatic
erythrocytes in bone marrow of mice as a
function of time after treatment with DMBA
or Glutathione, and DMBA pretreated with
Glutathione.

- 242 -

5. Abnormal Spermhead Assay Using Fumaric Acid against DMBA,
BaP, MNNG and MMC-induced Germ Cell Mutations

Experimental male mice of B6C3F1 strain were fed for one
week with powdered food containing 1% fumaric acid (FA) . On
the 8th day the mice were treated with mutagens in individual
groups. The doses were DMBA (20mg/kg; 2x) , BaP (20mg/kg), MNNC
(50mg/kg) and MMC (0.8mg/kg; 3x) . On the 35th day after last
injection, sperm smears were prepared and scored for abnormal
sperms. The results are presented in Table 5.

The results are similar to those obtained with fumaric
acid and the above mutagens in the micronucleus assay. In the
case of MMC, treatment with fumaric acid in the feed of mice
caused a reduction of about 45% in the abnormal spermliead count
However, fumaric acid had no significant effect in the case of
treatments with DMBA and BaP.

- 243 -

Table 5: Effect of Potential Inhibitors on Activity
of Mutagens using In Vivo Assays with Mice

Inhibitor

Bone Marrow Micronucleus Abnormal Spermhead

Mutagen Inhibitory
Effect

Fumaric acid

DMBA none DMBA

MNNG none BaP

MMC 35% inhibition MMC

none

minor inhibition

about 45% inhi-
bition

Glutathione

DMBA

60%

BaP

55%

3-carotene

DMBA

40%

BaP

50%

slight

MMC

variable (not
reproducible

- 244 -

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2. Bjelke, E. (1975) Int. J. Cancer 15: 561-565.

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21. Mitchell, J.R., D.J. Jollow, W.Z. Potter, J.R. Gillette, and
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27. Raj, A.S. and M. Katz (1984) Mutation Res. 136: 81-84.

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Sifter (1983) Fed. Am. Soc. Exp. Biol. Fed. Proc. 42; 2891.

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Mutation Res. 74: 347-356.

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35. Slaga, T.J. and W.M. Bracken (1977) Cancer Res. 37; 1631-1635.

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Forsch. 159: 85-91.

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"Carcinogens", Raven Press, New York, pp. 299-316.

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Res. 49: 2820-2823.

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H. Yagi, J.M. Sayer, D.M. Jerina and A.H. Conney (1982)
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42. Wyrobek, A. J. , and W.R. Bruce (1975) Proc. Natl. Acad. Sci.
(U.D.S.) 72: 4425-4429.

43. Wyrobek, A.J. and W.R. Bruce (1978) The induction of sperm-shape
abnormalities in mice and humans. In: A Holleander and F.J. de
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Their Detection", Vol. 5, Plenum Press, New York, pp. 257-285.

44. Wyrobek, A.J., L.A. Gordon, J.G. Burkhart, M.W. Francis,

R.W. Kapp, Jr., G. Letz, H.V. Mailing, J.C. Topham and M.W.
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- 247 -

Acknowledgements

This research was supported, in part, by a grant from
Provincial Lottery as Project No. 81-055-33 of the Ontario
Ministry of the Environment.

We acknowledge, with thanks, the technical assistance
of Mr. Anthony Wilson and Mr. Earl Stuart in the biological
assays.

We also wish to thank Shelton Dias, Ph.D., for his
valuable assistance in the chemical section of this research

- 249 -

Collaborative Study on Short-Term Tests

for Genotoxicity and Carcinogenicity

II - Carcinogen Assessment.

D.M. Logan

Department of Biology, York University,

North York, Ontario M3J 1P3

and

M.F, Salamone

Ontario Ministry of the Environment,

Biohazards Unit, P.O. 213,

Resources Road, Rexdale,

Ontario M9W 5L1

- 250 -

Col laboratlve Study on Short-Term Tests For
Genotoxicity and Carcinogenicity

D.M. LOGAN

Department of Biology, York University

For the last two years we have participated in an international
genotoxicity and carcinogenicity study established by The WHO, UN and
International Labour Office. The aim of this study is to develop a battery
of short term tests by which chemicals may be assessed quickly and at
relatively low cost for potential genotoxic hazard. The chemicals which we
have used (Benz (a) pyrene, pyrene, 2-acety lamino Fluorene and 4-acety lami no
Fluorene) are ones for which extensive whole life data Is available for
comparison. The assays tested were two in vivo assays, the bone marrow
micronucleus test and the abnormal spermhead test. In addition we have
tested the effect of two different routes of administration on the sensitivity
and response of the two assays, i.e. Intraperitoneal injection and gavage:
Benz (a) pyrene, a known carcinogen gives a positive reaction in both assays
while pyrene does not. 2AAF another known carcinogen also gives a p6sitive
reaction In both assays while the isomeric 4AAF does not. These data should
be contrasted with bacterial tests in which the noncarci nogens nevertheless
produce mutations and hence are scored as carcinogen positive. These assays,
particularly in common, therefore appears to offer a higher degree of
selectivity than that available in bacterial assays.

Seven additional chemicals including the two pesticides Mirex and
Atroclne which are of particular Interest to the Ontario Ministry of The
Environment were also tested and these data will also be presented.

- 251 -

INTRODUCTION

During the last three years laboratories in several countries have participated
In a collaborative genotoxicity testing programme. This is the second such
programme and is sponsored by The World Health Organization (WHO), The United
Nations Environment Programme and The International Labour Office. More
details on the first programme and the background of the second are included in a
previous publication (1). Briefly, the aim of the current programme is to identify
and characterise a group (or "battery") of tests by which genotoxic hazard may be
assessed. Such tests should be of relatively short duration (hence "short-term"), of
nnodest cost and offer high selectivity and reproducibility. The issue of selectivity
is particularly important. Although there exist a large number of biological tests
for genotoxicity no single test is completely selective. By this is meant a test
which identifies known genotoxins with 100% efficiency but non genotoxins are
always excluded, i.e. the test does not produce false positives. By using several
assays however it is hoped that their combined results will allow unequivocal
identification of genotoxins while excluding nongenotoxins. This is the basic
premise of the collaborative programme.

Our participation has involved the testing of four assay systems using known
carcinogens and non carcinogens. The four assays are; the in vivo mammalian bone
marrow micronuleus test, the jn vivo mammalian abnormal sperm head assay, the
in vitro sister chromatid exchange assay and the replicative/kiliing assay. In
addition a group of chemicals provided by the Ontario Ministry of the Environment
(MOE) were tested in the two in vivo assays.

- 252 -

Materials and Methods

Chemicals

The known carcinogens and non carcinogens were provided by Imperial
Cliemical Industries and in each case all the laboratories testing a particular
chemical received samples from a single synthesis to ensure comparability. AH
chemicals provided by The MOE were obtained from commercial suppliers in the
highest purity available.

Animals

The B6C3F1 hybrid mouse was used in both m vivo assays. All mice were
purchased from Canadian Breeding Farm and Laboratories Ltd., Quebec. Mice
used in a particular experiment were always selected from a single shipment and
were age matched in all cases.

Chinese Hamster Ovary (CHO) cells were kindly supplied by Dr. Richard
Marshall of Mutatech Inc.

Assays

The experimental protocols, selection of dose, dose sequence, sample times
and data scoring for each of the assays is presented in detail in the first paper of
this series (1). During the course of the study an additional assessment procedure
was proposed by The WHO. This involved evaluation of the WHO chemicals in the
same in vivo assays but using a different route of administration, i.e. gavage rather
than I. P. injection.

- 253 -

Results & Discussion

In Vivo Assays
(1) Abnormal Sperm Assay

(a) WHO Chemicals

Four chemicals provided by The WHO w/ere tested for genotoxic activity in
this assay. The chemicals v/ere two known carcinogens benzo (a) pyrene and 2 -
acetylaminofluorene (2AAF) and two non-carcinogens of related or isomeric
structure, pyrene and 4-acetylaminofIuorene C4AAF). The data obtained are shown
in Tables 1 and 2. In each case a control assay in which the known carcinogen 7,12-
dimethylbenzanthracene was tested provided a strong positive result. The dose
range reported was selected on the basis of previously reported toxicity data (1)
such that the maximum dose was 60-80% of the LD^g. The results are given as the
percentage abnormal sperm plus or minus the standard error. In the case of pyrene
and benzo (a) pyrene (Table 1) the assessment is straightforward. Regardless of
dose pyrene does not produce an increase in the rate of spontaneous abnormal
sperm production. In contrast benzo (a) pyrene shows a relatively smooth and
increasing rate of abnormality as the dose is increased and even by the second
lowest dose the rate of production is clearly above that found in the control (zero
dose). From this we assess pyrene as a negative germ cell genotoxin and benzo (a)
pyrene as a positive germ cell genotoxin.

The data for 2 AAF and 4 AAF (Table 2) are equally clear, i.e. over the dose
range tested 2 AAF was clearly genotoxic and 4AAF was not. It should be noted
that the dose range over which 4AAF was tested was lower than that with 2AAF.
This derives primarily from the chemicals' differing toxicities (LD^g).

- 2S4 -

(b) MOE Chemicals
Six chemicats provided by The Ministry of the Environment were tested in the
abnormal sperm assay. In the case of trichloroethylene, the assay was performed
using both the pure chemical and the pure chemical supplemented by the supplier
with a stabilizer which inhibits the chemicals breakdown or conversion to other
ciiemicals. The data obtained with four of the six are shown in Table 3. (Results
nf tests with mirex and atrazine have not been completed.) With the exception of
one treatment there is no evidence of genotoxic activity with any of the chemicals
tested. The possible exception is the value obtained with the highest dose of
trichloroethylene (minus stabilizer). In this case however, only one mouse survived
and the assay is currently being repeated. The lower doses with this chemical show
no evidence of mutagenic activity. The high dose (700 mg/kg) assay may therefore
be simply anomalous.

(2) Micronucleus Assay
(a) WHO Chemicals
The same V/HO chemicals tested in the abnormal sperm assay were tested
again using the micronucleus assay and the data obtained are shown in Tables 4 and
5. The data obtained indicate that 4AAF is not clastogenic while 2AAF is
clastoqenic. Note however with the 2AAF data that the sampling (assay) time
which gives the maximum number of micronuclei changes as the dose is changed.
Thus at the lowest dose only the earliest assay point is clearly above the control
while at the middle doses the intermediate sampling times produce the maximum
response. Finally at the highest dose the earliest sample time again shows the
most dramatic effect of the chemical and the percentage of micronuclei at later
sample: times is reduced. This variability reflects the complex interplay between
erythropoeisis, an animals response to a toxic chemical and the selection of
appropriate testing tirfies.

- 255 -

The data with pyrene and benzo (a) pyrene (Table 5) are also quite clear
although only a single drug dose was used in the case of pyrene. Pyrene is
apparently not clastogenic while benzo (a) pyrene is. However, the limited number
of assay points with pyrene indicates the need for more extensive confimation
assays.

(b) MOE Chemicals

The chemicals provided by the MOE were also assessed for their clastogenic
activity in the micronucleus assay and representative data are presented in Table
6. Chlorobenzene, 1,2-dichloroethane and pentachlorophenol are all clearly non
clastogenic (in each case the negative assay has been confirmed).
Trichloroethylene with or without stabilizer and Mirex are also apparently non
clastogenic and this is currently being confirmed. In the case of atrazine the
interpretation is less certain. Our initital results suggest a positive clastogen
response but these data are based upon a small number of surviving mice.
Confirmation experiments have also been troubled by toxicity problems and further
confirmation is underway.

The negative results obtained with the two in vivo assays must of course be
viewed with some reservations. A substance producing a confirmed positive result
in any in vivo assay must be considered a serious genotoxic hazard. The converse
unfortunately is much less certain and a negative response even if confirmed must
be accepted only within the limitations of the assay. In general, a substance which
fails to induce a positive response and is therefore assessed genotoxin negative is
probably not genotoxic. However many genotoxins are tissue specific, and may
require transport or metabolic supplementation which is foreign to the test
organism. In such cases the test chemical may appear inactive because of
metabolic "isolation" or failure to reach a specific target tissue such as bone

- 256 -

rrifirrow or gonads. For these and similar reasons a substance tentatively assessed
a;; non gonoLoxic can nevf3r be categorically so identified.
Route of Administration

As noted earlier, during the course of this study the WHO asked us to test the
effect of a different route of administration on the results of our ]n vivo assays. In
these cases the test substance was administered by gavage needle directly into the
stomach, rather by intraperitoneal Injection. Table 7 shows the results of one such
test in which pyrene and benzo (a) pyrene were tested in the micronucleus assay.
These data should be compared with those shown in Table 5. Again pyrene is
clearly non-clastogenic, while benzo (a) pyrene is a positive clastogen. These data
show quite clearly the importance of different assay times which was commented
upon above. In all case the earlier sampling time (47 hr.) shows a higher
clastogenic response than the later time (70 hr.). In fact at the higher two doses at
least the later sample time might be scored as clastogen negative.

In Vitro Assays
In both In vitro assays, the replicative killing and sister chromatid exchange
assays, all chemicals were tested both with and without metabolic activation
enzymes. These enzymes are thought to be essential for the conversion of
potential genotoxins into their active form. In our experiments they were prepared
from rat liver and since their preparation involves centrifugation at 9,000 xg they
are generally called an "S-9" preparation.

(1) Replicative Killing Assay

Assays involving the replicative killing assay were extended to include a wide
range of chemicals in addition to those assessed above. The data obtained are
summarised in Table 8. Of the positive i.e. mutagenic, chemicals three were
mutagenic in the presence but not the absence of an S-9 fraction. These chemicals

- 257 -

were 2-Aminoanthracene, 2-Acetylaminofluorene and Benzo (a) pyrene.

As was indicated by the in vivo assays 2AAF is mutagenic while AAAF is not.
In this case however activation was required to obtain a positive mutagenicity test.
Benzo (a) pyrene is also mutagenic but again requires activation. Of the chemicals
not assessed as mutagenic by this assay some have been identified as mutagens or
carcinogens in other systems, i.e. acrylonitrile (2), hexamethylphosphoramide (2)
and o-toluidine (2).

(2) Sister Chromatid Exchange (5CE) Assay
As with the Replicative Killing assay additional chemicals were tested in the
SCE Assay. Representative data are presented in Table 9. The positive control
substances produced a strong positive response but none of the test chemicals did
either in the presence or absence of activation factor (S-9). As noted above some
of these chemicals have been assessed mutagen positive in other assays. It should
be noted however that most of these chemicals were selected by the WHO because
they are generally not easily detected in most in vitro assays.

SUMMARY
A basic concept of the collaborative programme has been to test in various
different assay systems the genotoxicity of chemicals for which there exists a
large body of whole life and other corroborating data. From the numerous assays
tested it is then hoped to select a small group which taken together offer a highly
reliable indication of genotoxicity and more particularly carcinogenicity. We have
tested two in vivo and two in vitro assays with several suspected carcinogens and
non carcinogens. In the in vivo assays the chemicals tested may be paired i.e. one
is carcinogenic and the other is not. These paired chemicals were Benzo (a)
pyrene/pyrene, and 2AAF/4AAF, The first is carcinogenic, the second is not in
each pair.

- 258 -

Fioth in vivo assays are highly selective offering a clear positive assay of the
carcinogens and a negative assessnnent of the non carcinogens. One problem which
arises in the micronucleus assay deserves to be highlighted. That is the issue of
sampling time. The kinetics of erythropoiesis are complex and may be distorted by
the test chemical. Hence, any micronucleus assay must include a broad range of
sample times to ensure that the time of increased micronucleus production is not
missed entirely.

The route of administration does not seem critical in the assays we tested.
Both the effective dose range and the identification of clastogenic activity were
essentially the same for gavage and IP drug administration.

In contrast the data with the in vitro assays should be of concern. While
potent carcinogens produce a positive response there are two problems. First in
several cases a positive response occurs only when the chemical is supplemented
with a metabolic activator. The potentially complex interaction between a
ciiemical and activator(s) (which is not completely understood) introduces an
additional interpretative problem. For example, if neither chemical nor activator
is genotoxic but the combination is from which does the active principle derive i.e.
the chemical as normally assumed or the activator? Are all the ingredients of the
activator present in, for example, human tissue at the appropriate concentration?
It is clear that the need for activation (whatever the term implies) simply
complicates the interpretation.

The second and even more important concern with the in vitro assays is that
several suspected or known carcinogens do not produce a positive assay result over
the dose ranges tested. This data has been confirmed and indicates that at least
with the chemicals we have tested neither test offers the discrimination that is
being looked for. Since we do not yet know the results other laboratories have
obtained with these assays using other chemicals it may be that our selection of
test chemicals was a particularly unfortunate one. In the absence of such data

- 259 -

however the in vitro assays must be considered unsatisfactorily selective.

These data and extensive confirmation data will be submitted to the WHO
during 1985 and it is hoped that by the time of next year's conference substantial
progress will have been achieved in developing the desired test protocol for
carcinogens.

REFERENCES

.1. Logan, D.M. and Salamone, M.F. in The Proceedings of Technology Transfer
Conference //4 pt.I (1983, p. 283-300 - Toronto ISBN 6-7743-8797-1.

2. Sittig, M. in Handbook of Toxic and Hazardous Chemicals, Noyes Publications,
Park Ridge N.J. 1981, ISBN 0-8155-0841-7.

- 260 -

TABLE 1

Abnormal Sperm Assays of WHO Test Chemicals

Chemical

Total Drug Dose % Abnormal Sperm Comment

(mg/kg)

Pyrene

(5,0)*

255 (5,5)

510 (5,5)

770 (5,5)

1030 (7,5)

1285 (8,5)

1542 (10,5)

2.0 +0.3

1.8 ;o.2

2.2 +0.2

2.0 +0.2

1.8 ;o.2

1.9 ;o.2

1.9 +0.2

Assessed
Mutagen
Negative

Benzo (a) pyrene

(4,0)
138 (6,3)
210 (6,3)
279 (8,3)

348 (10,3)

420 (10,3)

1.7 +0.3

1.7 ;0.2

2.7 +0.5

3.2 ;i.4

2.9 ;0.6

4.6 +1.3

Assessed
Mutagen
Positive

* The bracketed figures are (Number of animals treated, number of treatments)

- 261 -

TABLE 2

Abnormal Sperm Assays of WHO Test Chemicals

Chemical

Total Drug Dose % Abnormal Sperm Comment

(mg/kg)

2-Acetylamino
fluorene
(2AAF)

(2,G)*

1100 (8,5)

2200 (8,5)

3300 (8,5)

4400 (8,5)

5500 (10,5)

6600 (10,5)

1.9 +0.2

2.0 +0.2
2.6 +0.3

4.1 +0.8
3.9 +0.7
4.1 +0.6
6,4 +0.9

Assessed
Mutagen
Positive

4-Acetylamino
fluorene
(4AAF)

(4,0)

180 (7,5)

365 (8,5)

545 (9,5)

735 (10,5)

910 (10,5)

1.9 ;o.2

1.3 +0.1

1.8 +0.2

1.5 +0.2

2.0 +0.2

1.5 +0.2

Assessed
Mutagen
Negative

* The bracketed figures are (Number of animals treated, number of treatments).

- 262 -

TABLE 3

Abnormal Sperm Assays of MOE Test Chemicals

Chemical

Trichloroetylene
(including stabilizer)

Trichoroethylene
(minus stabilizer)

Chlorobenzene

Pentachlorophenol

1,2-Dichlorethane

Total Druq Dose

LJrug
g7k5)

(5,4)*
100 (8,4)
148 (8,4)
200 (8,4)
250 (10,4)
300 (10,4)

84
104
128
160
200

30
60
95
115
155

(5,4)

(6,4)

(8,4)

(10,4)

(8,5)

(6,5)

(9,5)

(10,5)

(5,5)
46.5 (8,5)
70 (8,5)
92.5 (8,5)
130 (8,5)
140 (11,5)

125
250
375
500
625

(4,5)

(8,5)

(10,5)

% Abnormal Sperm

Comment

2.1

+0.5

2.0

+0.4

Assessed

2.1

+0.7

Mutagen

2.6

Negative

1.9

+0.7

1.9

+0.1

1.9

+0.2

2.6

+0.4

Uncertain

2.2

TO.l

because

1.6

+0.2

of high

1.8

T0.2

dose

4.4

2.1

+0.2

Assessed

1.9

+0.3

Mutagen

1.9

+0.3

Negative

2.1

T0.3

2.0

+0.3

No survivors

2.2

+0.3

2.1

+0.3

Assessed

1.7

+0.2

Mutagen

1.9

+0.3

Negative

1.5

+0.2

1.6

+0.2

1.8

+0.4

2.1

+0.3

Assessed

2.3

+0.4

Mutagen

1.7

?0.2

Negative

1.7

+0.3

1.6

+ .02

*The bracketed figures are (Number of animals treated, number of treatments).

- 263 -

TABLE 4
Micronucleus Assays of WHO Test Chemicals

Chemical

Drug Dose
mg/kg

Assay Time
(hours post
treatment)

Micronuclei
per 500 PCE

Comment

2-Acetylamino

0.5

+0.5

fluorene

220 (18,2)*

1.5

+0.6

0.3

+0.2

0,5

+0.5

Assessed

660 (18,2)

1.3

+0.6

Clastogen

1.7

+0.9

Positive

1.0

+0.5

1110 (24,2)

1.2

+0.4

2.7

+1.1

1,2

+0.6

15A0 (24,2)

6.0

+2.5

2,3

+1.2

3.0

+0.4

A-Acetylamino

€

0.25 +0.25

fluorene

346 (36,2)

0.5

+0.2

Assessed

0.6

;o.2

Clastogen

546 (24,2)

47
70

0.5

;o.2

Negative

* The bracketed figures are (Number of animals treated, number of treatments).

- 264 -

TABLE 5

Micronucleus

Assays of WHO Test Chemicals

Chemical

Drug Dose
mg/kg

Assay Time
(hours post
treatment)

Micronuclei
per 500 PCE

Comment

Pyrene

(6,2)*

1.2 +0.4

Assessed

822 (24,2)

1.3 +0.6

Clastogen

0.8 +0.4

Negative

0.4 +0.2

Benzo (a)

(6,2)

0.2 +0.2

pyrene

44 (18,2)

2.7 +0.9

Assessed

2.7 +0.6

Clastogen

2.2 +0.6

Positive

88 (15,2)

48
69
80

3.6 Tl.l
4.0 T0.4

1.7 +0.6

133 (20,2)

48
69
80

4.0 +0.7
6.3 +1.4
3.3 +0.5

177 (19,2)

48
69
80

2.2 +0.5
4.0 +0.8
4.7 +0.8

* The bracketed figures are (Number of animals treated, number of treatments).

- 265 -

TABLE 6

Micronucleus

Assays of MOE Test Chemical

Chemical

Drug Dose

Assay Time

Micron uclei

Comment

mg/kg

(hours post
treatment)

per

500 PCE

Chiorobenzene

(6,0)*

1.3

;0.6

25 (18,1)

2.0

+0.3

1.2

+0.3

Assessed

0.4

+0.2

50 (18,1)

1.2

+0.4

Clastogen

1.0

+0.5

2.2

+0.4

Negative

114 (24,2)

0.6

+0.4

0.4

+0.3

100

1.0

+0.6

Dichloroethane

(6,0)

0.5

+0.3

100 (18,1)

0.3

+0.3

Assessed

0.7

+0.5

0.7

+0.3

Clastogen

200 (18,1)

1.2

+0.6

Negative

1.2

^0.2

Atrazine

Inter-

(3,0)

0.7

+0.3

pretation

331 (15,1)

2.5

+1.0

Uncertain

364 (15,1)

1.0

due to low

400 (16,1)

4.0

+ 3.0

survival

Mirex

(4,0)

1.0

+0.6

600 (17,2)

1.4

+0.8

Assessed

0.8

+0.2

Clastogen

600 (15,1)

0.9

tO.4

Negative

0.5

T^O.4

cont..../2

- 266 -

TABLE 6 (cont.)

Chemical

Drug Dose

Assay Time

Micronuclei

Comment

mg/kg

(hours post
treatment)

per 500 PCE

Pentachloro-

(5,0)

0.2 +0.2

phenol

32 (10,2)

1.5 1^1.5

Assessed

37 (10,2)

1.0 ^1.0

1.0 +1.0

Clastogen

42 (10,2)

0.2 +0.2

Negative

46 (15,2)

48
72

0.6 ?0.2
1.0 +0.4

Trichloro-

(5,2)

0.2 +0.2

ethylene

50 (8,2)

0.25 +0.2

(plus stabilizer)

0.25 +0.2

Assessed

60 (8,2)

0.25 +0.2

Clastogen

70 (8,2)

0.5 +0.5

0.5 :f-0.5

Negative

80 (12,2)

48
1%

0.2 +0.2
0.2 +0.2

Trichloro-

(6,0)

0.7 :f0.3

Assess-

ethylene

ment

(minus

40 (8,2)

0.7 +0.4

Uncertain

stabilizer)

due

60 (10,2)

to low

Survival

80 (12,2)

The bracketed figures are (Number of animals treated, number of treatments).

- 267 -

TABLE 7

Micronucleus Assays of Pyrene and Benzo (a) Pyrene

Test Drug Administered by Gavage

Chemical

Drug Dose
mg/kg

Assay Time
(hours post
treatment)

Micronuclei
per 500 PCE

Comment

Pyrene

(7,2)*

0.6 +0.3

311 (14,2)

0.3 +0.2

Assessed

0.2 +0.2

Clastogen

466 (14,2)

47
70

0.1 +0.1

Negative

700 (14,2)

Benzo (a)

(7,2)

0.6 +0.3

pyrene

120 (14,2)

1.6 +0.6

1.1 +0.3

Assessed

150 (14,2)

2.6 +0.7

Clastogen

Q.4 +0.2

Positive

187 (14,2)

47
70

3.0 +0.9
0.9 +0.3

* The bracketed figures are (Number of animals treated, number of treatments).

- 268 -

table:

Repllcative Killing Assay of Selected Chemicals

Chemical Mutagenicity

N-Methyl-N-nitro-N-nitrosoguanidine +

2 Nitrofluorene

2 Aminoanthracene + ^*

2 Acetyaminofluorene - * ^

4 Acetyaminofluorene

Benzo (a) pyrene * 3

O-toluidine

Benzoin

Caprolactam

Safrole

Hexamethylphosphoramide

Phenobarbital
Die thy list ilbestrol
Acrylonitrile

* The suffix a after a positive indication of mutagenicity indicates that a
positive result was achieved only when the test chemical was supplemented

ith an S-9 activation mixture.

- 269 -

TABLE 9

Sister Chromatid Exchange Assays of Selected Chemicals

Chemical Mutagenicity

Dimethylnitrosamine (control) +

Methylmethane sulfonate (control) +

Hexamethylphosphoramide

O-toluidine

Caprolactam

Diethylstilbestrol

Safrole

All the test chemicals which scored negative were tested both with and
without added activation factors (S-9).

- 271 -

THE DEVELOPMENT OF A FRESHWATER FISH TEST
TO IDENTIFY AQUATIC TOXIC CONTAMINANTS

I.R. Smith and V.E. Valll, 1984

Department of Pathology,

Ontario Veterinary College,

University of Guelph,

Guelph, Ontario, NIG 2W1.

ABSTRACT

Approaches to assess mutagenicity and/or carcinogenicity In all
fish species are necessary to facilitate routine monitoring and trend
analysis of genotoxlc Inputs. The suitability of fish embryos and
juveniles, and of techniques for detecting chromosome damage, were
determined utilizing Brachydanlo rerlo (Zebrafish) as a model species-
Previously published results (Smith and Valll, 1983) were encouraging,
and further investigations are reported here.

The acute effects of Ethyl methanesulphonate (EMS) Included terato-
genesis after both 24 hour and 8 day embryo exposures, which could be
related to tissue (cell) death. In addition, latent effects evident
after short-term exposure periods included mortality and inherited muta-
tions In epidermal cells (leading to cell death through apoptosls). A
latent effect not induced by EMS (after 6 months observation) indicates
that the genotoxicity observed In embryos did not lead to neoplasia.

Further genotoxicity analysis revealed that newly fertilized
embryos were more sensitive than 6 hour old embryos, and that erythro-
cytic micronuclel (30 hours of age) were a more sensitive endpoint than
yolk-sac micronuclel. The most sensitive measure of genotoxicity yet
examined (24 hour anaphase aberration analysis) Indicates a sensitivity
equivalent to published results utilizing adult fish and more complex
techniques .

The sensitivity and wide applicability of fish embryos (plus the
relatively low cost) make the examination of either laboratory or feral
fish, exposed in the laboratory or in situ , a promising technique. The
examination of temporal and spatial variability of genotoxicity end-
points in wild fish embryos as they relate to Industrial discharges may
be the most promising application. Further applications Include assess*
ing the relative Impacts of parental chemical contributions, or water
uptake, and the further study of latent effects, including neoplasia and
mutations.

- 272 -

INTRODUCTION

The widely huld suspicion chat 80-90% of human cancers are related
to environmental factors (Berenblum, 1974) has led to a search for the
identity of those factors. The search for environmental causes of
cancer has focused primarily on food and occupational sources of expo-
sure. The possible impact of Industrial waste on rivers and lakes used
as a waste repository and of these discharged chemicals on man (via
drinking water» recreational use, watering of livestock or crops » or
fish consumption) or aquatic organisms has received less attention.

The presence of carcinogenic chemicals in the aquatic environment
has been reviewed and discussed (Kraybill, 1976, 1977; Kraybill, Helmes
and Slgman, 1978; Borneff, 1977; Allnk, 1982). It has been shown that a
variety of water bodies contain carcinogens at detectable levels,
Kraybill et al^. (1978) found 64 suspected or proven carcinogens, and 27
promotersTcocarcinogens out of the 1728 chemicals found. In addition,
65 of these 1728 chemicals were found to be mutagens, including 27 for
which no carcinogenicity data was available. The role of water as a
repository for carcinogenic chemicals Is receiving Increasingly more
attention, as evidenced by a survey (Allnk, 1982) which found 862 papers
published between 1977 and 1982 concerning aquatic carcinogens. Sources
of Industrially discharged carcinogens and mutagens have been
identified. These include pulp and paper mills (Nestmann, Lee, Mueller
and Douglas, 1979; Douglas et^ al • , 1980; 1982; Kinae, Hashizuma. Makita,
Tomica. Klmura and Kanomorl, 1981a, 1981b), and coke and coal ovens (van
Talcott, 1979; Moore, Osborne and Davies, 1980; Osborne, Davles, Dixon
and Moore, 1982) has also been shown to contain these chemicals.

Two basic approaches exist for detecting these hazardous
contaminants. Chemical analysis of concentrated water samples has
detected many carcinogenic and mutagenic chemicals. Confirmation of the
biological activity of these chemicals has been by testing the pure
chemicals using laboratory animals- The second type of approach
measures an appropriate response in wild organisms as an indication of
both the hazardous chemical's presence, and Its measurable activity.
Subsequent determination of the chemical that Is responsible is
performed on extracts of the organisms which detected Che chemical's
effects initially. This approach has become available because of the
development of a variety of short- term tests for genetic damage
(genotoxiclty) Indicative of somatic mutations. The somatic mutation
theory of cancer has developed from the finding that most mutagenic
compounds are also carcinogenic (Ames, Durstan, Yamasakl and Lee, 1973;
Miller, 1978; McCann £t al^. , 1979). Short-term approaches utilizing
testing for mutagenic compounds have been a widely utilized route for
the detection of potential carcinogens.

The possible implications of waterborne carcinogens/mutagens are
many. Epidemiological assessments of cities have found a limited
correlation between polluted drinking water and cancer (see the review
by Allnk, 1982). Treated drinking water has been shown to be mutagenic,
reflecting prior contamination and the effects of treatment (Kool, van

- 273 -

Kreljl, van Hansen and DeGeef , 1981; Alink, 1982). The impact of
mutagens /carcinogens on fish includes the induction of both tumours and
mutations which have been detected in the laboratory and in wild fish.
Tumor induction may result in mortality, and mutation induction in
offspring would conceivably lead to reduced recruitment, because most
heritable mutations are lethal. Brown, Hazdra, Keith, Greenspan,
Kwapinski and Beamer, (1973) found that fish with a higher tumour
Incidence also had a higher incidence of non-oncogenic disease, perhaps
Indicating additional immune system effects due to the carcinogenic and
other chemicals present. Mans consumption of fish, which have been
shown to accumulate mutagens from contaminated waters , poses an unknown
threat .

The monitoring of chemically induced changes in aquatic
populations, as indicators of both immediate and long-term effects,
removes the reliance on chemical concentration procedures (which attempt
to mimic biological accumulation of contaminants) necessary for chemical
analysis or laboratory testing with bacteria or mammals. Recently, many
authors (see the reviews by Kligerman, 1982a, 1982b; Landolt and Kocan,
1983) have Indicated that fish exposed to mutagens are affected in a
variety of ways, many exhibiting various types of chromosome damage and
breakage. These studies have employed a limited number of fish species,
with limited applicability for both the field and laboratory. To
monitor effects In the environment, and test effluent components under
controlled laboratory conditions, a test system should be adaptable to
any fish species, and be widely applicable to a variety of waters and
laboratories. By utilizing a life-stage characterized by rapid cellular
division, fish testing may be more sensitive to both short and long-
terra effects. Simple methods of analysis for chromosome damage and/or
breakage would make analysis quick and Inexpensive, without sacrificing
sensitivity or accuracy for detecting somatic genotoxlclty.

The analysis of fish eggs from chemically Impacted areas or
contaminated parents should provide an integration of contaminant levels
and types. Chemical bloconcentratlon In adults and passage to the
embryo via yolk transfer (Hose, Hannah, Landolt, Miller, Felton and
Iwaoka, 1981), gamete effects (including mutations), and chemical uptake
during fertilization and development from the water are all potential
exposure routes for embryos. Embryos from uncontaminated areas or
laboratory animals would facilitate the analysis of Individual chemicals
or Inputs, via chemical uptake from the water, or gamete/parent exposure
through either food or water. The advantages of utilizing this approach
Include a high division rate (leading to a high sensitivity to
genotoxlcants), a multitude of acceptable species, and ease of exposure
and maintenance. The long-term consequences of the genetic damage
Inflicted on the embryos can also be assessed, Including cancer
Induction, teratogenesis , altered fecundity and reproductive success,
mutagenicity, etc . , all measures which are difficult to assess In fish
exposed as adults.

Indications of genetic damage are detectable through several
possible techniques. The high mitotic rate of the embryo should supply

274 -

a large number of anaphases-telophases for analysis of chromosome
damage. Evidence of cell death could also be observed, due either to
genetic damage or cytotoxicity. The detection of micronuclel, becoming
popular In mammalian work (Heddle, Hite, Kirkhart, Mavournin, MacGregor,
Newell and Salamone, 1983) is also a possibility, possibly reflecting a
cumulative summation of genetic damage, as micronuclel may be
persistent. The developmental rate of the organism could be monitored
to assess whether or not Inhibition of development was present. These
types of damage were monitored in embryos exposed to EMS and MIK to
assess the described indicators of genotoxic damage, and the possibility
of differentiating genotoxlclty from cytotoxicity, eliminating concern
about false positive responses. Toxic effects including neoplasia
induction, acute lethality and extended testing of embryo-larval stages
provided further evidence of the effects of the test chemicals.

- 275 -

MATERIALS AND METHODS

Approaches for procuring and exposing embryos and fry have been
published previously (Smith and Valli , 1983; Smith, 1984). Details will
be confined to experimental design's and specific approaches.
Several tests were used to assess the acute and subacute toxic effects
of both compounds on fry and embryos. A second series of tests was
designed to assess Che genotoxlclty of the test compounds with embryonic
zebraflsh. The proven mutagen and carcinogen EMS was tested as a
representative "positive" mutagen because of demonstrated genotoxlclty
In other aquatic organisms. MIK was chosen as a negative control
because of a structural similarity to EMS.

Lethal Testing: Ethyl methanesulphonate

Two 96 hours bioassays were performed, one with zebraflsh
juveniles, and one with zebraflsh embryos. Two replicates of 30
juvenile zebraflsh (6 weeks old, weighing less than 10 mg each) were
exposed to concentrations of 0.0. 62.5, 125, 250, 500 and 1000 mg/L EMS
In 250 ml dechlorlnated water, with the solutions being completely
renewed dally. Dissolved oxygen levels were measured after 24 hours and
48 hours, and pH after 48 hours utilizing a Rexnord Model 650
Multianalyzer . The 96 hour LC50 for embryos was determined with 3
replicates of 30 embryos (8 hours old at initiation of exposure) exposed
to 30 mL of test solution containing 0, 31, 62, 125, 250, 500 mg/L EMS,
solutions being renewed dally. Eggs from 3 females were pooled prior to
selection, to obtain sufficient embryos.

Embryo-larval testing: Ethyl methanesulphonate

The determination of effects on embryo and larval stages over 8
days was performed on three replicates of 30 organisms each, obtained
from the pooled eggs of 3 females . Eight hour old embryos were exposed
to 30 mL teat solutions containing 0, 31, 62, 125, 250, 500 mg/L EMS, at
25 +/- l^C; solutions were renewed dally. Survival and hatching rates
were determined dally, and surviving teratogenic organisms were counted
and examined on day eight. Changes in body shape and structure were
assessed and described Including scoliosis and kyphosis, under 40X
magnification. All replicates were combined after day 8 and placed in
control water for continued observation until day 10.

Delayed effects testing: Ethyl methanesulphonate

Repeated juvenile exposure

Fish were exposed on days 1, 6, 12 and 15 to 900 mL's of test
solution for 24 hours without aeration, and removed to control water
between exposures. EMS levels were 0, 0.1, 1.0, 10, 100 and 1000 mg/L.
150 juvenile (3 weeks post-hatching) fish weighing 6 mg each were
exposed to each concentration, at a temperature of 23*'C. Mortalities
were recorded dally for the initial 2 weeks of exposure, and survival
determined on days 16, 39, 80 and 130. Between exposures the test

- 276 -

organisms were held In 8 L tanks at 25°C in the fume hood, and fed brine
shrimp nauplii. Dissolved oxygen levels and pH readings in the EMS
solutions were measured after the final exposure period.

On day 21, the fish were transferred to static 50 L glass tanks
with aerated ZS'^C control water. On day 80, the fish were transferred
to flow-through 100 L tanks receiving 1 L/rain aS^C dilution water,
gradually being switched to a diet of frozen adult brine shrimp.
6 months after exposure to EMS male/female pairs from the control, 100,
and 1000 mg/L groups were spawned. Values for the number of spawns,
total number of eggs/spawn and survival of 100 randomly chosen eggs to
24 hours of age were recorded. Tumour frequencies were determined in
the control and 1000 mg/L groups at this time, and other pathological
changes were noted.

Single Embryo Exposures

Delayed effects due to EMS were examined with an exposure regime
equivalent to that of the genotoxicity measurements. Fish were exposed
to the test solution for 24 hours (beginning as blastulas), after which
they were held in control water for observation. Embryos were obtained
from the pooled eggs of 3 females. Five replicates of 30 eggs each were
exposed to 0, 0.1, 1.0, 10, 100 and 1000 mg/L EMS, in a total volume of
30 mL. After exposure the embryos were transferred to 100 ml beakers,
containing 30 raL water. Survival and hatching rates were recorded
daily, with teratogenesls monitoring on days 4, 5, 6 and 8. Feeding
with Pararaecia commenced after 90% hatching, and the fish were kept with
a 24 hour photoperiod to facilitate feeding. 50% of the water volume
was replaced daily after day 8 and the fish were transferred to 200 mL
water in 250 mL beakers on day 25, and then to 50 L static and
eventually 100 L flow- through tanks as described previously.

Non-mutagen toxicity

MIK toxicity was determined in two ways. A 24 hour lethality
bioassay with newly hatched juvenile zebrafish, 5 per replicate, used 20
raL of test solution, containing 4000, 2000, 1000, 500. 250, 100 and
mg/L MIK, Test solutions were not renewed. The 4-day embryo-larval
assay began with 6 hour old embryos exposed to 2000, 1000, 500, 100, 10,
I and mg/L MIK, mimicking the concentration range of the EMS
exposures. Solutions (20 mL per replicate) were changed daily in all
three replicates, of 30 organisms each. Survival, mortality and
hatchability were monitored dally, and teratogenesls was determined at
96 hours .

GENOTOXICITY TESTING

The protocols have been previously described for this testing
(Smith and VaUi, 1983), the only deviation being that exposure of
embryos began at 8 hours of age, rather than 2 hours, as previously
described. Analysis of anaphase aberrations and erythrocytic
micronuclei was undertaken only after 24 hours of exposure.

- 277 -

RESULTS

Results from the first portion dealing with some genotoxlclty
aspects of this study have been published previously (Smith and Valll,
1983).

Acute Lethality: Ethyl methanesulphonate

The 96 hour bloassay with juveniles produced mortalities only In
the 1000 mg/L group» which exhibited 100% mortality. Consequently, the
96 hour juvenile LC50 was graphically (problt scaled paper) estimated as
700 mg/L. The embryo bloassay yielded partial mortalities in the 500
mg/L concentrations of less than 50%, which was insufficient mortality
to estimate an LC50. Lethality data from the embryo-larval tests first
four days produced an LC50 of approximately 700 mg/L,

Water quality was unaffected by EMS, no conslstant concentration
dependency being present In any values, though lethal levels did have
lower dissolved oxygen levels (3.5 mg/L) .

Embryo-larval Testing: EMS

The embryo-larval assay resulted in an MATC substantially below the
LC50 previously determined, teratogenesls being the most sensitive
indication of EMS toxicity. A summary of the results obtained with this
test is contained in Table I. Survival on day 8 was significantly
concentration dependent (ANOVA). The lowest concentration significantly
different from control was 250 mg/L, which was also significantly
different from all other concentrations (T-test). The mortality evident
by day 10 in both the 250 and 125 concentrations after they were removed
to clean water may reflect some latent effect.

Teratogenesls frequencies were significantly related to
concentration (ANOVA) both as absolute frequencies and when adjusted for
survival. Only the values In the 125 and 250 mg/L concentrations
consisting of 3 or more abnormal Individuals could be considered a
significant increase above control (T-test). The types of defects
present Included scoliosis in 31 of the organisms, lordosis in 2
organisms, and an elongated/extended yolk-sac in 3 organisms. Scoliosis
was most common in lower concentrations, while the full variety of
changes was present In 250 mg/L.

Significant variation In hatching with concentration was present on
all days (ANOVA), and on day 5 both the 125 and 250 mg/L concentrations
had a significantly higher hatching rate than control, perhaps
reflecting minor temperature differences. Control temperatures were
24''C, 500 mg/L temperatures 25*'C. On day 6 the 250 mg/L hatching rate
was significantly below the 125, 62 and 31 levels, the absence of a
difference with control possibly being temperature related. Day 7 and 8
found a lower level in 250 mg/L than all other concentrations. The 250
mg/L concentration thus reduced hatching success on days 6, 7 and 8.

- 278 -

Table I: Embryo- larval testing results for an 8 day exposure of

Brachydanlo rerlo to Ethyl raethanesulphonate- Values are
means (x/30) for three replicates of 30 organisms each,
unless otherwise stated, for absolute responses, hatch-
ability and teratogenesis not being corrected for survival
(s.d.).

Test
Determinant

Time
(day)

Concentration (mg/L)
62 125

250

500

Survival

Successful
Hatch

Terato-
genesis

5
6
7
8

28.6(0.6) 28.3(1.1) 29.0(2.1) 29.3(1.1) 12.6(5.5) 0.0(0)

78 82 74 2

1.3(0.6) 3.0(2.6) 5.0(3.6) 8.6(2.5) 6.0(1.7) 0.0^

U.3(3.8) 15.6(4.9) 21.0(6.1) 26.0(2.6) 6.0(1.7) 0.0°

27.3(0.6) 25.0(1.0) 29.0(1.0) 30.0(0.0) 8.6(4.2) 0.0**

27.3(0.6) 27.3(2.1) 29.0(1.0) 30.0(0.0) 10.0(5.2) 0.0**

8 0.0(0.0) 0.3(0.6) 0.6(1.2) 8.3(0.6) 2.6(0.6)

c _

^ Combined replicates, this value being the number of survivors out of
the original 90 test organisms, which were pooled on day 8.

^ Some embryos hatched, however they died immediately after hatching.

^ This value is strongly influenced by mortality, being 20% of survivors,
the value for 125 mg/L being 28%.

- 279 -

Additional observations made during this test Included a lack of
pigmentation in the 500 mg/L exposed embryos on day 2, Indicating a
reduced developmental rate. After combining all replicates on day 8,
the fry resulting from 250 and 125 mg/L EMS remained on the bottom of
the vessel and didn't feed. In contrast to the remainder of the
organisms. The MATC for EMS encompasses the range of 62-125 mg/L based
on teratogenesls-

Delayed Effects: Ethyl methanesulphonate

Both exposure regimes resulted In little acute lethality ,al though
teratogenesls was caused by embryo exposure for 24 hours to 1000 mg/L,
and delayed mortality was induced in both groups. No apparent
differences between survival In controls and 1000 mg/L were evident in
the fish exposed to repeated 24 hour exposures (Table 2) on day 16. A
significant decline in survival was evident in the 1000 mg/L
concentration by day 38, survival being only 17% (control was 73%).
This Indicates that considerable delayed toxicity was Induced, which was
expressed after day 16. Of the mortalities occurring up to day 39 In
the 1000 mg/L group, 65% were post-exposure, and 42% occurred in a
single one week period between days 20 and 27, 4-11 days post-exposure.

Water quality during the last 24 hour exposure period was monitored
In the test vessels , and dissolved oxygen levels (mg/L) ranged from 4 .0
In the 1000 mg/L to 5,0 in the 0.1 mg/L vessels, levels low enough to
possibly have contributed to some of the mortalities observed during
exposure. The pH was unaffected by EMS, ranging from 6.5 in control to
6.9 in the highest concentrations.

During the holding (130 days) of test organisms, the 1000 mg/L
survivors tended to grow faster, due perhaps to a lower number of fish/
tank, however they did not feed well, were easily startled, and swam
less vigorously than other survivors. The pattern of colouration of the
1000 and to a lesser extent the 100 mg/L exposed fish was altered, the
straight yellow lines on the control fish being replaced with broken,
often wavy patterns in the higher concentration fish. Additionally,
several large masses (2-3 mm In diameter) were observed deep within the
dorsal musculature of several 1000 mg/L fish when they were held up to a
strong light.

Histological examination of 10 fish from the control, 100 and the
1000 mg/L concentrations revealed no tumors in any fish. The masses
seen grossly were areas of Internal hemorrhaging, due perhaps to damage
during netting, or other procedures.

Large numbers of anaplastic epidermal cells were present in the
fish previously exposed to 1000 mg/L EMS. Many karyorrhexic and
pyknotlc cells were also evident in the epidermis of these fish, in a
multifocal distribution. Few such cells were present in control fish.
These findings may have been related to the alterations in body
colouration and pattern observed in exposed fish.

- 280 -

Table 2: The survival of Brachydanlo rerlo after 4 exposures to

various concentrations of Ethyl raethanesulphonate on days
1, 6, 12 and 15. Survival is expressed as the number
remaining out of the original 150 organisms.

Day

Concentration (mg/L)
0.1 1 10

100

1000

111

109

105

130

105

82 83
69 74
68 73

107
26
23
23

Table 3

Endpoint

The effects of a single 24 hour exposure of embryos to
various concentrations of Ethyl raethanesulphonate, removed
to dilution water_after exposure. All values are means for
30 organisms (x/30) each for 5 replicates except where
indicated (s.d.)

Survival

Terato-
genesls

Time
(day)

0.1

Concentration (mg/L)
1.0 10.0

100.0

1000.0

9 29.0(1.2) 29.8(1.2) 28.8(2.7) 29.6(0.5) 27.4(1.8) 28.8(0.8)

19 20.0(3.8) 22.0(2.3) 20.8(4,6) 18.0(3.3) 3.6(4.2) 0.0(0.0)

110^ 85 89 89 17

7 0.0(0.0) 0.4(0.9) 0.0(0.0) 0.0(0.0) 0.2(0.4) 25.0(1.4)

^ Combined replicates, n=150 initially (day 0).

- 281 -

Breeding results were widely variable, and no significant
differences were evident between the control and 100 mg/L fish. Four
pairs from the 1000 mg/L group did not spawn although they were allowed
a total of 14 days. The control fish (control male and female) spawned
on 15 of 26 possible days, with an average total production of 117
eggs/spawn, average infertile or dead when siphoned from the tank bottom
of 2.9% and average survival at 24 hours of 76.5%. The 100 mg/L fish
spawned 6 of a possible 12 days, with an average total egg production of
97/spawn, a slightly higher 7.8% dead at spawning, and a slightly lower
survival of 63.7% at 24 hours. The absence of spawning in the 1000 mg/L
fish is unexplained, for the fish examined appeared to be normal
histologically. The ratio of males to females was determined for 25
fish from most concentrations (23 from 1000 mg/L) by visual means, to
vary from 1.5:1 to 1.2:1 in control, 0.1, 1, 10 and 100 mg/L groups,
however the ratio was 4.75:1 for the 1000 mg/L group indicating a
doubling of the relative number of male fish.

A single 24 hour exposure of embryos to EMS resulted in
teratogenesis and delayed mortality similar to that described
previously. The survival of the exposed embryos on day 9 ranged from
91% to 99%, control being 96.6%, indicating no effects of prior
treatment on survival. Beginning on day 9 however both the 1000 and 100
rag/L concentrations (Table 3) induced significantly greater (T-test)
mortalities than in controls. Survival on day 19 when the replicates
were combined averaged 66, 73, 69, 60, 18 and 0% in controls, 0.1, 1,
10, 100 and 1000 mg/L respectively, a significant reduction with
concentration (ANOVA). Limited mortalities were present after day 20,
and by day 110, survival rates were 57, 59, 59, 51, 11 and 0%
respectively for 0, 0.1, 1, 10 100 and 1000 mg/L.

Significant teratogenesis was limited to the 1000 mg/L group,
containing an average of 87% abnormal fish (of survivors). The
abnormalities were divided roughly Into two groups. Lordosis/kyphosis
(75 fish) a..d scoliosis (6 fish) accounted for 65% of the abnormalities,
while the remainder (35% or 44 fish) featured a "helical peduncle", in
which the body didn't straighten after hatching, being curled with a
longitudinal twist. Many of these individuals also featured a shortened
body. On day 7, the 0.1 mg/L group had one fish with scoliosis and one
fish with a distended abdomen. The 100 mg/L concentration produced a
single fish with a helical peduncle.

Exposure of various zebrafish life stages to EMS resulted in
considerable delayed mortality, and teratogenesis when exposed as
embryos. No tumors were induced; however breeding success was affected
in the highest concentration groups and an unusual sex ratio (4-74:1
male: female) was present.

Non-mutagen Toxicity

The 24 hour bloassay of MIK with juvenile zebrafish yielded 100%
mortality In 1000 mg/L, while 500 mg/L was completely non-lethal,
yielding a graphically estimated LC50 of 700 mg/L. The 4-day embryo-

- 282 -

larval test produced few effects below 1000 mg/L. Survival due to 1000
mg/L EMS was significantly lower than control on days 2, 3 and 4 (T-
test). Hatching was unaffected with the exception of 2000 mg/L (which
was zero). All organisms hatched by day 5, after removal to control
water. The frequencies of teratogenesls on day 4 were 0, 2.3, 0, 3.5,
14.8 and 80.7% as a percentage of hatched survivors, for control, 1, 10,
100, 500 and 1000 mg/L MIK, respectively. All non-zero values were
statistically different (T-test) however the absence of an increase in
the 10 mg/L group means that only 100 mg/L and above induced significant
teratogenesls. Deformities were predominantly scoliosis, with a few
dwarfs and fish with lordosis. The fish with scoliosis included many
which resembled the EMS Induced helicle peduncle, but without
longitudinal twisting. A no-effect level, based on the 24 hour
lethality and 4-day embryo-larval tests with MIK was 10 mg/L, the MATC
being 10-100 mg/L.

Anaphase Abnormalities: Ethyl methanesulphonate

The control AA rate at 24 hours was 0.88/20 (n-9, S.D.-0.78)
Insignificantly different from that in earlier experiments. Exposure of
embryos 6 hours older than prevloulsy utilized (Table 4) EMS Induced
significant variability with concentration (ANOVA) though only 100 and
1000 mg/L Induced a significant Increase above control (T-test). The
effect of a 6 hour delay on the sensitivity of the embryos is evidenced
by the lower abnormality frequency in this experiment than previously in
both 1000 and 100 mg/L EMS (T-test, p < 0.05), and at a reduced
confidence Interval (P < 0.1) in 10 and 1 mg/L as well, indicating
reduced sensitivity. Overall, 100 mg/L EMS Induced a significant
Increase In AA levels in embryos exposed Immediately after
fertilization, after 6 and 12 hours exposure, or in embryos exposed as
blastulas, when sampled at 24 hours. The most sensitive protocol
utilized recently fertilized embryos sampled after 24 hours of exposure,
revealing a significant increase due to 1 rag/L EMS.

Generally, the types of damage present In controls were present in
all EMS doses, though an induction particularly of acentric and attached
fragments was evident, many of which were scored as lagging fragments.
After 24 hours exposure, acentric and attached fragments were induced
with Increasing concentration as were multiple defects.

Mlcronuclel: Ethyl methanesulphonate

Erythrocytic tnlcronuclei (MN) were induced in a concentration
dependent fashion after 24 hours of exposure to EMS. Erythrocytic
mlcronuclel were normally closely associated with the nucleus, and their
size ranged between l/20th and 1/lOth of the nuclear diameter. The
nucleus was often Indented, the mlcronuclel being associated with the
Indentation.

Erythrocytic mlcronuclel levels were less variable than those in
yolk-sac cells (Table 4). A significant induction above control by 10
mg/L EMS and a concentration dependency were found for erythrocytic

- 283 -

Table 4a: Anaphase aberration rates in embryos exposed to EMS for 24 hours.

Concentration (mg/L)
0.1 1.0 10 100 1000

0.88 0.80 0.90 1.20 3.60 7.80

(0.78; 9) (1.32; 10) (0,57; 10) (1.03; 10) (1.58; 10) (1.87; 10)

Table 4b: Mlcronuclei levels in erythrocytes of embryos exposed for 24 hours
as in Table 4a.

Concentration (mg/L)
O.l 1 10 100 1000

0.05 (0.09) 0.07 (0.11) 0.14 (0.16) 0.16 (0.17) 2.09 (0.35) 5.21 (2.85)

- 284 -

mlcronuclet. EMS had no significant effect on the number of
erythrocytes available for analysis. Erythrocytic micronucleatlon rates
versus anaphase aberration rates show a similar significant
concentration dependency as that seen in Experiment A, indicating that
the same or similar factors result In both anaphase aberrations and
raicronuclei .

Erythrocytic mlcronuclei appear to be more reliable indicators of
genotoxicity, exhibiting greater sensitivity and less variability than
yolk-sac cells. A larger erythrocytic population is available for
analysis, and the precise time of formation (beginning at 26 hours at
25*0) is known.

Fyknosls and Karyorrhexis: Ethyl methanesulphonate

Pyknotlc and karyorrhexlc cells were assessed utilizing a
semi-quantltatlve assessment of damage and a significant induction of
cell death was evident. Control organisms were assigned an average
value of O.ll, while values of 0.2. 0.45, 0.35. 3.6 and 4,9 were
assigned to 0.1. 1. 10. 100 and 1000 mg/L EMS respectively. Significant
Increases were evident In 100 and 1000 mg/L EMS (Mann-Whitney test). In
a similar fashion as in Experiment A both 100 and 1000 mg/L EMS induced
considerably larger amounts of damage than other exposed groups. It is
significant that while the 100 and 1000 mg/L groups lived for 9 days,
delayed mortality was evident in these concentrations.

Non-mutagen Effects

Anaphase aberrations were not Induced in 24 hour old embryos by MIK
when analyzed at levels up to concentrations leading to widespread
cellular degeneration. AA levels (x/20 (S.D.)) of 1-2 (0.79), 0.25
(0.5), and 0.89 (0.8) were present in 0. 500 and 1000 mg/L respectively.
No tested values were significantly higher than control (T-test). The
mitotic rate was so low In embryos exposed to 2000 mg/L MIK that
insufficient anaphases were available for analysis (the embryos were
moribund) .

The influence of MIK on micronucleatlon rates in rbc ' s was
pronounced. The average control rate (x/100) was 0.06 (S.D. = 0.09),
while in 500 and 1000 mg/L levels of 0.31 (0.32) and 0.42 (0.39) were
significantly higher than those of control (T-test, ANOVA) . The
induction of mlcronuclei by MIK may be related to cell death. In all
EMS treated groups with high AA rates and MN rates and In 500 and 1000
rag/L MIK treated groups, a large proportion of the cells were dead.
Fyknosls/ karyorrhexis relative values of 0.0, 0.8 and 2.05 for control,
500 and 1000 mg/L MIK exposed embryos were evident, only the 1000 mg/L
value being significantly higher than control (Mann-Whitney). 2000 mg/L
MIK led to very high numbers of pyknotics qualitatively. Fragments from
karryohexlc nuclei strongly resemble mlcronuclei, and their presence
within the cytoplasm of the erythrocytes (possibly phagocytozed, or
accumulated during the formation of cytoplasm) may have Influenced these
findings.

- 285 -

MIK at a level of 1000 mg/L significantly (T-test) reduced the
mitotic Index of the exposed embryos from the control value of 42.1
(S.D. = 7.4), to a level of 29.5 (6.8). 500 mg/L MIK had no effect
(46.8 (9.5)), while 2000 mg/L MIK resulted In very few mitosis.

- 286 -

DISCUSS [ON

Acute Effects: Ethyl methanesulphonate

EMS was relatively non-lethal and provided little Indication of
being chronically toxic, in that its MATC was only 10% of its acutely
lethal level. The suspected volatility and rapid degradation of this
chemical possibly combined to reduce its lethality, and In this respect
EMS may be similar to many discharged volatile direct-acting mutagens.
EMS produced teratogenesis at levels below those causing lethality,
showing sublethal effects at the tissue level. The teratogenesis
induced by continuous exposure could also be induced with exposure only
during the first 24 hours of life. The production of teratogenesis did
show that EMS can alter tissue development, possibly by reacting with
DNA, leading to cell death and heritable defects.

Low acute and sub-acute lethality of EMS was also found in
published studies with adult fish. EMS up to 200 mg/L was non-lethal to
adult raudmlnnows and killiflsh after 2 days of exposure (Hooftman. 1981;
Hooftman and Vink, 1981). Prolonged exposure (6 weeks) to 200 mg/L EMS
led to 100% mortality in adult mudmlnnows (Hooftman and de Raat, 1982).
Further evidence for the low toxicity of EMS has been reported by
Samoiloff £1 ^l- (1980), who found no effects in a long-term
developmental test with nematodes up to levels of 124 mg/L, close to the
no-effect level in this study.

The teratogenesis Induced after exposure to EMS is indicative of
tissue/growth alteration prior to or during early organogenesis.
Alteration of DNA by EMS could exert a very general effect on developing
tissues, including muscle. Other possible causes could be the general
alkylating properties of EMS, altering proteins and other macromolecules
necessary for proper development.

Extensive studies on BaP have produced teratogenesis in gastrulas
(Hose e^al., 1982) and reduced hatchability at 0.1 ug/L BaP. Two ug/L
BaP has been shown to cause mortality in rainbow trout embryos (Hannah
et al., 1980), 0.21 ug/L BaP resulting in teratogenesis at hatching.
TeraTogenesls has been shown in a Cyprinodontidae exposed to dibutyl
phthalate (mutagenic and carcinogenic) at 10% of the juvenile LC50
(Koenig e^ al . , 1982). BaP, dibutyl phthalate and EMS all cause
abnormalities at approximately 10% of the lethal level. It appears that
mutagenic chemicals in general are not potent teratogens, and the
induction of teratogenesis by a wide variety of pollutants. Including
metals, organics, pesticides and physical factors (Laale, 1981;
Bengtsson. 1975; Sloof, 1982) indicates that this endpolnt is not of
diagnostic value for mutagens in particular.

Non-mutagen Toxicity

The Induction of teratogenesis by the negative control chemical
(MIK) at 500 rag/L (70% of the LC50) indicates that this chemical, can
produce teratogenesis. A wide variety of metals, organic chemicals.

- 287 -

pesticides and physical factors also can produce teratogenesis. The
acute lethality of MIK was minor, the only finding of interest being the
teratogenesis induced.

Delayed Effects: Ethyl methanesulphonate

Several effects of EMS were noteworthy; the considerable latent
mortality resulting from both embryo and juvenile exposures, the altered
sex ratios, coloration changes, and reduced reproductive success evident
6 months after exposure of the juvenile organisms. These effects are
indicative of the induction, In a very short period, of time of very
persistant changes.

The lag in mortality (4-9 days) after exposure indicates that some
vital process was affected which was essential during later life.
Possibilities include the disruption of energy pathways, or of other
processes requiring large amounts of gene expression, including RNA
synthesis and erythropoests. Delayed mortality beginning 4 days after
the final juvenile exposure and 9 days after the embryo exposure
resulted in up to 100% mortality. Mortality at hatching was reported
for embryos exposed to 0.5 mg/L of the pro-rautagen Aflatoxin Bl (AfBl)
for 2 hours (Wales, Slnnhuber, Hendricks, Nixon and Elsels, 1978).
Similarly, trout embryos exposed to 30 and 100 mg/L of the direct-acting
mutagen MNNG for 1 hour (Hendricks, Wales, Slnnhuber, Nixon, Loveland
and Scanlon, 1980) had reduced growth during the following year,
possibly Indicating latent effects. The higher mortality in embryos
exposed to 1000 mg/L and 100 mg/L EMS for 24 hours, versus the 4 x 24
hour exposure of juvenile fish may reflect a higher sensitivity of
embryos to the latent effects, the LCSO's for these two ages being
similar. The exposure of trout embryos to MNNG (Hendricks et al . , 1980)
failed to produce such effects. Sensitivity varying with age has been
shown to AfBl effects (cumulative 90 day mortality) after a single one
hour exposure to 0.5 mg/L, showing that older embryos and juveniles were
quite a bit more sensitive to the mutagenicity of this chemical
(Hendricks £1 al • > 1980). This was felt to be due to the Increase in
hepatic mixed function oxidase levels In older fish. AfBl must be
metabolized to form a mutagenic Intermediate, whereas EMS is probably
metabolized to a nonmutagenic Intermediate. EMS may be more toxic to
earlier embryos, due to the lower levels of hepatic metabolism, than in
juveniles, the liver not forming until 24 hours of age.

The abnormal sex ratio in the 1000 mg/L exposed fish may have been
due to sex related mortality, or may have been chemically Induced by the
action of EMS on pregonlal cells. These have been shown to be present
as early as the 16-1000 cell stage (Walker and Strelsinger, 1983).
Grossly, no secondary morphologic sex characteristics in normal
zebrafish exist beyond the dlsteaded belly of the mature female and
slight coloration differences, but histological examination of 10 fish
confirmed that no "males" were actually Immature females. A similar
finding, that trout embryos exposed to MNNG (high concentrations)
produced a preponderance of male fish (Hendricks e£ al . , 1980), may also
have reflect delayed effects or hormonal Influences.

- 288 -

The abnormal coloration and epidermal karyorrhexls may be related,
as the epidermis Ln the fish exposed as juveniles would have been
intimately exposed to the direct acting mutagen. The epidermal and
pigment cells may have been mutated by the EMS, leading to changes in
coloration due to abnormal cell development, cell death, or altered
migration patterns. If mutations were Induced in epidermal cells,
subsequent divisions may have resulted In cell death of one, both, or
neither of the daughter cells, depending on the Inheritance of the
suspected lethal characteristic (dominant or recessive). This type of
phenomenon may be responsible for both the abnormal pigmentation, and
the observed epidermal cell death 6 months after exposure.
Karyorrhexls is associated both with necrosis and apoptosis (Wyllle,
Kerr and Currie, 1980). Apoptosis, including karyorrhexls, normally
occurs in diffusely scattered cells. In the fish examined in this
study, multifocal areas contained scattered areas of karyorrhexic cells,
often in large numbers. It has been hypothesized that apoptosis can be
due to genetic damage, leading to cell death (Wyllie et^ a]^- , 1980). The
detection of nuclear aberrations, including karyorrhexic cells, has been
suggested as a test procedure for detecting the effects of mutagenic
chemicals in tissues with a high mitotic rate (Dr. M. Goldberg, Ontario
Veterinary College, University of Guelph, Guelph, Ontario., Pers.
Comm.). Such tissues include epidermis, kidney, intestinal epithelium
and kidney, although in younger fish, growth related mitosis would be
considerable in all other tissues. The absence of changes in tissues
other than skin Is unexplained, though it may be related to the higher
concentration at the skin during immersion.

The reproductive failure of the fish exposed to 1000 mg/L is
unexplained, given the apparently normal histological character of the
testes and ovaries in the fish examined. It is possible that hormonal
alterations resulting in the changes in the sex ratio may have also
affected some spawning processes. The lack of a reproductive failure in
100 rag/L indicates the limited effects of EMS treatment, in only
affecting reproduction at very high concentrations.

Similar juvenile exposures to various carcinogens (Schultz and
Schultz, 1982a, 1982b) have resulted in increased tumor incidences. The
high delayed mortality in 1000 mg/L EMS may have contributed to the lack
of tumors, as only 23/150 fish survived.

Anaphase Aberrations

EMS clearly induced significant levels of chromosome damage,
despite some variability in control levels. The tested embryos
exhibited sensitivity to a very wide range of EMS concentrations up to
LOOO mg/L. This level, which is ultimately lethal, caused abnormalities
in 75% of the observed anaphases (Smith, 1984; Smith and Valll, 1983),
in contrast to an average control level of 4%. Chromosome damage was
readily observed after only 6 hours of exposure to low levels of EMS,
and a significant increase above control was evident in 100 mg/L at 6
and 12 hours, while only 1 rag/L was necessary to cause a significant
increase after 24 hours. The sensitivity of embryos varied with the age

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at exposure Initiation, as well as with exposure duration. Embryos
exposed initially as early blastulas responded significantly only to a
level of EMS 100 times that having a similar significant effect when
embryos were exposed as 16-64 cells . All exposures initiated with
blastulas resulted In 40 to 52% fewer damaged anaphases than when
exposure began with the earlier stages.

Background (control) AA rates of approximately 4% in this study are
comparable to those ranging from 1,5% to 6.5% reported by other authors
for a variety of fish species and ages (Hose et al., 1982; Pechkurenkov,
1973; Tsoy, 1974; Longwell and Hughes, 1980). The low control rates
observed for fish embryos apparently extend to sea urchin embryos (Hose
et al., 1983; Hose and Puffer, 1983). The high mitotic rate of the
embryo may be a factor in this rate not being zero. The background rate
in a fish cell line (12%) was considerably higher (Kocan e£ al • , 1982)
possibly due to its transformed state.

Criticism of the use of AA's for mutagen detection and testing has
concentrated on the higher levels of control damage with this technique
(Kligerman, 1982a; 1982b). A comparison of AA control rates can be made
with metaphase analysis by Isolating acentric and attached fragments
which correspond to chromosome or chromatic breaks and gaps respectively
at metaphase. Published results for organisms with a lower mitotic rate
than embryos (adults) generally find gaps and breaks in to 0.5% of the
cells (Kligerman e£ al^., 1975; Hooftman, 1981; Hooftman and Vlnk, 1981;
Krlshnaja et al., 1982), though one report found breaks in 8% of gill
cells (Prien et^ a]^. , 1978). Acentric fragment control levels in this
study were 0.2%, while the attached fragment frequency was 1.6%,
slightly higher than most metaphase studies. Sea urchin embryos control
levels of these defects were 0.3% and 0.8% (Hose e£ al . , 1983; Hose and
Puffer, 1983), while In a cell culture, control levels of acentric and
attached fragments were 4.6 and 3.6% respectively (Kocan et^ al. , 1982).
The similar to somewhat higher rates of control defects of these types
may reflect an increased sensitivity in discerning small scale breakage
or gaps rather than the induction of artifacts. An additional advantage
of AA analysis is the visualization of events recognized only with
difficulty at metaphase. Lagging chromosomes and bridges are thought to
result from assymetrical interchanges and induced "stickiness", possibly
as a result of chromosome breakage (Nichols et^ al . , 1984) however other
events may also be responsible (Gaulden, 1982), which don't have a
complementary metaphase abnormality, including spindle fiber
malfunctions and pre-metaphase alterations which can lead to aneuploidy
(Danford, 1984; Liang, Hsu and Henry, 1983). The observation that these
types of defects form the majority of the control lesions in this and
other studies (Hose and Puffer, 1983; Hose eit al^. , 1983; Kocan et al.,
1982) may have led to the belief of high background rates in AA
analysis, when in fact AA analysis may be able to detect events which
are not discernable at metaphase.

EMS Induced considerable nuuibers of aberrant anaphases over a 3
order of magnitude concentration range. Levels which, when administered
for the first 24 hours of life caused insignificant mortalities until 19

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days post-exposure, Induced damage in 75% of the anaphases at 24 hours
(Smith, 1984). This amount of damage appears to have had little
Immediate effect. It Is possible that abnormal anaphases couldn't
complete mitosis, and may have accumulated, artificially Increasing the
sensitivity of this technique. The finding of a significant increase
above control In 1 rag/L EMS (Smith and Valli, 1983) indicates a similar
sensitivity to approaches using longer periods of exposure with adult
fish and metaphase analysis (Hooftman, 1981; Hooftman and Vink, 1981),
and a sensitivity only slightly less than that reported for SCE
Induction by EMS (Alink et^ al^- , 1982) induced over a much longer time
period. The AA test approach has similarly been shown to be as
sensitive as SCE analysis or metaphase analysis when used in cell
cultures, or with sea urchin embryos for a variety of chemicals (Kocan
et al , 1982; Hose et al^. , 1983).

The induction of concentration dependent effects (plus the absence
of AA induction by MIK) up to a level affecting the majority of the
cells Indicates that a valid genotoxlc response is being measured. The
types of defects induced by EMs, (including acentric and attached
fragments which have been induced by EMS and monitored at metaphase) all
responded in a concentration dependent fashion, reaching levels well
above control In 100 mg/L after only 6 hours. The increased sensitivity
of the AA test appears to be that it includes types of damage not
visible at metaphase resulting from Induced "stickiness". These were
elevated above control levels at similar or lower levels than chromosome
breakage events. When combined with breakage events, this provides a
more accurate representation of the genotoxic events occurring.

While this study has shown AA analysis to be as sensitive as other
approaches to a direct acting genotoxic chemical, work with sea urchin
embryos (Hose e£ al . , 1983) has shown a similar sensitivity to the
pro-mutagen, BaP, as these other approaches. The role of embryonic
enzymes In activating BaP to a reactive intermediate may provide a clue
to the shifts in multiple damage seen after 12 hours (Smith, 1984). It
has been shown (Todd and Bloom, 1982) that EMS genotoxiclty decreased
after liver development, possibly as a result of detoxification.
Cytochrome P-450 and P-448 have been implicated in the activation of
pro-rautagens to mutagens as well as the detoxification of many chemicals
(see Parke, 1981). It has been shown that these enzymes are detectable
prior to liver formation in fish embryos (Binder and Stegraan, 1982)
indicating that extrahepatic tissues may play a substantial role in
chemical metabolism (Binder and Stegman, 1980). Changes in enzyme
activity, levels, or the formation of a rudimentary liver by the age of
24 hours may have been responsible for the reduction of multiple damage.
While in the higher concentrations of EMS (100 and 1000 mg/L) the number
of multiple defects dropped off, the total number of AA remained similar
or Increased, possibly indicating that the mechanism by which multiple
defects are formed is different from that of single defects. It may be
(Longwell, 1978: Longwell and Hughes, 1980) that bridges, lagging
chromosomes and multiple damage (mostly bridges) are due to induced
stickiness, rather than breakage and reattachment. The induction of
multiple "sticky" chromosomes by MNNG in grasshopper neuroblasts (Liang

- 291 -

and Gaulden, 1982; Gaulden, 1982) has been observed, which ultimately
lead to chromosome breakage , leaving fragments or the remainder of the
chromosome lagging. The reduction In multiple damage at 24 hours was
not paralleled by a reduction of single bridges or lagging chromosomes.
It Is possible that multiple damage is due to a multiple hit phenomenon,
or that a threshold must be exceeded before their Induction. Multiple
defects have been induced by other mutagens (Kocan e^ aJL . , 1982; Hose et_
al., 1983; Liang and Gaulden, 1982) and were not observed In experiments
with non-mutagens at inhibitory (Kocan £t al_. , 1982) or cyto-toxic (this
study) levels. The absence of Induction of any abnormalities by any of
these non-mutagens leads to the supposition that toxicity does not cause
AA. Multiple abnormalities at metaphase have been induced by mutagens
(Krishnaja and Rege, 1982; Kligertnan, 1975; Hooftman and Vink, 1981) so
It is not unlikely that some multiple bridges actually reflect multiple
chromosome damage as well as non-disjunction due to stickiness . The
possibility also exists that uneven or unequal distribution of EMS
within the embryos may have led to higher concentrations In some cells.
The multiple layers of cells present after gastrulatlon may have
restricted the penetration of EMS, which is relatively non-lipophilic,
Into the deeper layers of cells, reducing the possibility of multiple
damage types. The induction of multiple damage in embryos exposed
Initially as blastulas indicates that a reduction In chemical concen-
tration over the 24 hours of exposure is not responsible. Considerable
variability in the induction of multiple damage after 24 hours exposure
was evident, for an unexplained reason.

The induction in a concentration dependent fashion of AA's In
several test organisms by a wide variety of mutagens (Hose .^^ £l_- . 1983;
Kocan et al . , 1982; this study) leaves no doubt that this technique Is
appropriate for detecting genotoxic events In aquatic organisms. Some
of these events are visible at metaphase, however evidence to date
suggests that the assessment of damage at anaphase inducible by
phenoraenan other than breakage and the sensitivity of the embryo in
general (high mitotic rate) make this approach as sensitive as any other
types of assessment. Further elucidation of the phenomenon underlying
various types of defects (especially multiple defects) and the effect of
age at exposure initiation and sampling is warranted.

Micronuclei

The concentration dependent Induction of micronuclei in two cell
populations by EMS (after 24 hours exposure) resulted in a significant
induction above control levels of this type of damage in 10 mg/L EMS for
erythrocytic micronuclei and 100 mg/L for yolk-sac micronuclei (Smith
and Valli, 1983). The yolk-sac approach yielded higher levels of
control damage than erythrocytic control levels and the yolk-sac
micronuclei levels were much more variable, for both control and exposed
organisms, than the erythrocytic MN. Good correlations between AA and
MN rates Indicate that similar phenomenan are responsible for their
occurrence. The wide variability in damage estimates and the tedious
approach necessary to separate karyorrhexic nuclear fragments from
micronuclei reduce the value of this approach for yolk-sac cells. The

- 292 -

sensitivity uslag erythrocytes was excellent however. Erythrocytic MN
were a more sensitive indicator of genetic damage in organisms initially
exposed as blastulas than AA analysis In the same organisms. In view of
the apparently greater sensitivity of embryos exposed at the 16-64 cells
stage, the greater sensitivity of the erythrocytes approach is even more
noteworthy, as the yolk-sac MN were monitored in organisms initially
exposed in the 16-64 cell stage.

A disadvantage of the ralcronuclei approach for aquatic studies IS
the presence of nucleated erythrocytes, obscuring a portion of the
cell's volume, however MN have been detected in fish and amphibian
erythrocytes (Hooftraan and Raat , 1980; Siboulet, Grinfell, Deparis and
Jaylet, 1984). The use of yolk-sac cells has not been reported for MN
analysis, though it has been suggested for cytogenetic analysis
(Longwell, 1978). While the large cytoplasmic volume and size of the
yolk-sac cells was thought to optimize the visualization of MN, several
factors combined to make this approach of doubtful usefulness. The low
numbers available for analysis and the presence of karyorrhexic
fragments made analysis very tedious. Frequencies of MN were signifi-
cantly higher only at levels of EMS causing considerable pyknosis and
karyorrhexls, and the poor correlation with EMS levels and AA damage may
reflect a poor sensitivity to EMS. This may in part be due to the cell
cycle of the yolk-sac cells. They are still dividing at 24 hours in
some areas, while other areas feature static or regressing cells. It
has been shown that cells undergoing a natural death (common In embryos)
exhibit apoptosis (Wyllle, Kerr and Currie, 1980) which results in small
karyorrhexic-like nuclear fragments being phagocytosed by adjacent cells
in many tissues; yolk-sac cells may be subject to artifactual increases
in these micronuclei-like inclusions. The small number of cells
available for analysis in part reflects the difficulty in obtaining good
preparations of these cells without obscuring yolky material. This
statistical downfall may have contributed to the wide variability, as a
good representative sample of yolk-sac cells was difficult to obtain.

The measurement of erythrocytic ralcronuclei was less variable and
more sensitive than yolk-sac cells. Good concentration dependency and
correlation with AAs indicate that this type of damage may be less
susceptible to artifacts due to cell death. The larger number of cells
available for scoring make this approach more statistically viable.
Erythrocytes are first formed between 26 and 32 hours of age, and thus
are not subject to apoptosis, as are the yolk-sac cells. The absence of
phagocytosis in mature erythrocytes makes it less likely that necrotic/
apoptotic debris will be confused with ralcronuclei in the immature
erythrocytes. The erythrocytic micronucleus test has become very well
accepted (see Heddle et_ al . , 1983; 1984) and has been adapted to foetal
mouse erythroblasts (Cole, Cole, Henderson, Taylor, Arlett and Regan,
1983; Cole, Taylor, Cole and Arlett, 1981; Cole, Taylor, Cole, Henderson
and Arlett, 1982), the precursor cells of erythrocytes, and to
granulocyte-macrophage progenitor cells (Henderson, Cole, Cole, Cole,
Aghamoharamadl and Regan, 1984).

- 293 -

The detection of ralcronuclel in erythrocytes due to 10 mg/L EMS
shows a sensitivity much greater than that found for adult fish.
Studies have shown damage in mature erythrocytes which required up to 6
weeks exposure to 8 mg/L EMS to induce MN (Hooftman and Raat, 1982).
This may in part be a reflection of the cell cycle (rate of division) in
the mature animal. Work with newt larvae (Siboulet ejt al_. , 1984) after
hatching found that MN due to X-irradiation were present in the greatest
numbers 3 days after treatment » while mouse embryo erythrocytes showed a
maximal increase 48 hours after BaP treatment (Cole et_ al . , 1981). The
shortest cell cycle is probably present in Che youngest organisms, the
organisms in this study forming erythrocytes from very primitive
tissues, at a great rate. This, coupled with the proximity of liver
tissue (which produces the erythropoietic cells, would make this
approach very sensitive for pro-mutagens which require Qetabollsra.
Metabolism of direct acting mutagens by the liver may detoxify them
rather than activate them, hence the Increased activity apparent in
hepatic erythrocytes may reflect the high division rate, and a low
detoxifying capacity at this early time. It has been shown for foetal
mice (Cole et_ a]^. , 1983; Henderson et a^. , 1984) that the foetal
erythrocytic MN were a more sensitive indicator of chemical genotoxlcity
than the polychromatic erythrocytes from the parent's bone marrow, when
exposed to methyl methanesulphonate (direct acting) and other
chemicals -

Consideration of the source of these MN centers mainly on acentric
fragments which are not incorporated into the nucleus after division,
and possibly lagging chromosomes, due to spindle fibre malfunction or
damage (Evans, Neary and Williamson, 1959; Carrano and Heddle, 1973;
Heddle, 1973; Schmid, 1977). The finding in this study that both types
of micronuclei were correlated with the total number of defects at
anaphase, including acentric fragments and lagging chromosomes corrobo-
rates the supposition that damage at metaphase only becomes a micro-
nucleus if it results In lagging chromosomes or pieces of chromosomes,
visible at anaphase. The finding that micronuclei in erythrocytes were
a more sensitive indicator of genetic damage than AAs conflicts somewhat
with this, as only a portion of the damaged chromosomes at metaphase
form micronuclei (Evans e£ al_. , 1959; Carrano and Heddle, 1973). It is
conceivable that this is because some broken fragments are not Isolated
at the first anaphase, and thus wouldn't show up in AA analysis, but may
be expelled at later divisions, or at Interphase. The wide variability
in damage due to lower levels of EMS may be responsible for the reduced
sensitivity of AAs in a statistical manner, rather than the type of
damage being a less sensitive indicator. The assessment of all types of
AA damage may have increased the statistical variability, due to the
inclusion of types of damage which are invariably lethal to the cell
(multiple damage, multi-polar figures and possibly bridges) and wouldn't
result in MN.

The accumulation of cells with MN may also Increase its
sensitivity, providing a cumulative index of damage because the
erythrocytes do not divide, and hence many potentially lethal lesions
are not expressed. In this way, microuucleated erythrocytes can survive

- 294 -

and their numbers buildup, while damage seen at anaphase disappears
after mitosis Is completed. Variability in the division rate of the
embryo Itself may make AA analysis in most tissues less sensitive than
AA and MN analysis in erythropoietic tissue, which are dividing very
rapidly at this time. This is in contrast to the remainder of the
embryo, which while it has the same number of division figures per
embryo as at 6 hours, has a lower division rate. A similar situation is
thought to be present in the mouse embryo, contributing to the greater
foetal sensitivity over the parent tissues (Cole et^ al^- » 1983). The
analysis of additional organisms for AAs may take this reduced
sensitivity less of a problem, as it is more efficient to count AAs than
MN, facilitating the use of larger numbers of organisms. Sea urchins
exposed to BaP (Hose et^ al . , 1983; Hose and Puffer, 1983) showed that AA
analysis at gastrula was more sensitive than micronuclel in embryo
cells, paralleling the situation found in the present study. No studies
have as yet utilized early fish embryos, in which erythropoiesis is
occurring, and monitored both AAs and MN in the same organisms. Mouse
embryo erythrocytic micronuclel were as sensitive an indicator as SCE
analysis in the same tissue (Cole et^ al^. , 1983). SCE analysis is
generally regarded as more sensitive than AA or metaphase analysis
(Kligerman, 1983a; 1983b) in adult fish and other organisms. This may
infer that the foetal erythrocytic mlcronucleus test Is a very sensitive
assay, when compared to either investigations of chromosome damage, or
to studies of any damage with adults.

While yolk-sac micronuclel provided a relatively poor indication of
genotoxlclty due to EMS, erythrocytic micronuclel in 32 hour old
organisms were more sensitive than AA analysis In this age of organism,
but less sensitive than AA analysis in 26 hour old organisms, all after
24 hours of exposure. It would appear that the exposure of organisms
from the 16-64 cell stage with monitoring of AAs at 24 or 32 hours and
erythrocytic micronuclel at 32 hours, would be the most sensitive assay
for EMS. In general it appears that the damage visible at anaphase is
due to the same damage resulting in MN, and may in some cases be
responsible for it, given the good correlations between these types of
damage .

Pyknosis and Karyorrhexls

Induced cell death, as evidenced by pyknosis and karyorrhexls, was
considerably higher than control in 100 and 1000 mg/L EMS exposed
embryos. These concentrations also resulted In teratogenesls and
delayed mortality in acute effects testing. Elevated cell death in EMS
concentrations Inducing significant numbers of AAs and MN indicates that
the same mechanism Is responsible for all three types of damage, or
alternatively that AAs and MN result In cell death. Significant MN and
AA Induction was also encountered at mutagen levels which did not
produce appreciable cell death, indicating that cell death was not as
sensitive an indicator of genotoxlclty, nor that cell death was respon-
sible for either MN or AA.

- 295 -

Significant cytotoxicity due to EMS was restricted to
concentrations greater than those causing significant increases in
genotoxicity indicators, and was characterized by wide variability.
This may have been due in part to the heterogeneity of the embryo, as it
was common to observe isolated areas In which pyknosls and karyorrhexls
were abundant, making It difficult to obtain a representative index.

It has been suggested that anaphase aberrations lead to mlcronuclel
and/or cell death (Longwell, 1978; Kocan et^ al^. , 1982). For a
ralcronucleus to be present in one of the daughter cells, the other cell
must be short of a considerable amount of DNA, possibly even entire
chromosomes. Many of the types of damage seen at anaphase could lead to
cell death, including multiple defects, multipolar figures and possibly
bridging. These may delay or prevent the formation of nuclear and
cytoplasmic membranes, and undoubtedly at least one of the daughter
cells would be deficient In some DNA. The mechanisms underlying the
appear ance of AAs may be in part breakage, in combination with
"stickiness" derived phenomenon, as described earlier. The events
leading to stickiness, as well as those leading to breakage, could cause
cell death, either by a chromosome deficiency, or the inability to
complete mitosis. Reduced cell survival has been reported to be
correlated with SCE's (Tofllon, Williams and Deen, 1983; Krelger and
Garry, 1983; Morris, Heflich, Beranak and Kodell, 1982), mlcronuclel
(Heddle e£ al^. , 1983; Heddle and Salamone, 1981), and single strand
breaks (Loch-Caruso and Baxter, 1984). From the present study, AAs
appear to be closely related to cell death. Mlcronuclel are routinely
counted only at 50-80% of the LD50, because levels resulting in cell
death due to chromosome breakage must be reached. In the present study,
AAs were detected at levels as low as 1% of the ultimately lethal level,
or 0.15% of the LD50.

In organisms surviving for 9 days after exposure to 100 and 1000
mg/L EMS the correlation between AAs and cell death was variable. In
organisms wich 75% (Smith and Valli, 1983) abnormal anaphases, 84% of
the cells observed were dead, while levels causing 37% abnormal
anaphases had only 3.2% dead cells. The prevalence of delayed mortality
in 1000 mg/L EMS may be related to cell death. 100% delayed mortality
occurred due to 1000 mg/L, while 88% mortality occurred In 100 mg/L,
possibly also being related to cell death, which was significantly
higher in those groups. The variation in cell death between 100 and
1000 mg/L EMS couldn't be completely ascribed to any type of AA, with
the possible exception of multiple damage. The high levels of multiple
damage in 1000 mg/L at 12 hours in Experiment A or 24 hours in
Experiment B relative to 100 mg/L correlate well with the observed
difference In cell survival after 24 hours exposure. Only the multiple
damage rates at 24 hours in Experiment A fall to correlate with this.
It may have been that the sensitivity of the cells at this stage
resulted in their early death, without their being visible as anaphases
in sufficient numbers, or their development being delayed.

While it has been shown that mutagens can cause cell death In fish
embryos (Geraudle, 1981; Hannah, Hose, Landolt, Miller, Felton and

- 296 -

Iwaoka, 1982; Hose et_ aj^. , 1982; Hose and Puffer, 1983; Hose et al.,
1983) cyto-toxlcity Is by no means restricted to mutagenic or genotoxlc
substances. The alkylating properties of EMS probably result In the
alteration of a wide range of biological macromolecules . a process which
could contribute to cell death, without altering genetic materials. It
is apparent however that a substantial outcome of genetic damage Is cell
death. Recently some authors have drawn attention to this phenomenon as
a phenomenon (apoptosls) in dividing tissues (Goldberg, BlaJcey and
Bruce, 1983; Searle, Lawson, Abbott and Kerr, 1975; Wargovlc, Goldberg,
Newmark and Bruce, 1983) due to mutagens and antl-prollferatlve agents.
Embryos provide excellent populations of dividing cells for such
studies, but the background rates may be somewhat variable In later
embryos. Analysis at earlier stages of development may provide a better
correlation between AAs and cell death.

At 24 hours in this study, cell death was always a less sensitive
Indicator of genotoxiclty than AA, as sensitive as yolk-sac mlcronuclel,
but less sensitive than erythrocytic mlcronuclel. Studies with sea
urchin embryos (Hose and Puffer, 1983; Hose et^ al • » 1983) found mixed
results; in one study there was significant cell death at Che same
mutagen levels causing AAs, in the other finding cell death at higher
levels, while results with fish embryos (Hose e£ al^. , 1982) were not
quantified. The use of gastulas would be preferable, due to lower
control rates noted for sea urchin gastrulas. Control rates In the
early embryos used In this study were 0.7%. The control group
previously identified as having the lowest survival and highest control
AA and MN rates also had the highest rate of pyknosls/karyorrhexls.
This substantiates the role of genetic aberrations as possibly being
responsible, leading to increased cell death in late embryos, and
possibly to Increased embryo death at the sensitive gastrula stage.

The role of cytotoxicity (cell death) In the observed teratogenesls
seems plausable. 1000 mg/L EMS resulted In teratogenesls In 87% of the
survivors, while 100 mg/L resulted In only 1% teratogenesls. This may
reflect the wide difference in pyknotlc cell numbers previously noted
between these two concentrations (84% versus 3.2%). Given that the MN
and AA levels only varied by a factor of 2 (in both Experiments A and B)
between 1000 and 100 rag/L, the Increased teratogenesls seems to be more
closely related to cell death than genotoxiclty. This may also Indicate
that lower levels of AAs (for example in 100 mg/L) are repaired to some
degree, while 1000 rag/L exerts sufficient damage that repair was unlike-
ly, leading to cell death and teratogenesls. This has been observed In
teratogenic mouse limbs due to methyl nitrosourea (Loch-Caruso and
Baxter, 1984). Fish cell DNA repair after mutagen exposure has been
noted (Walton, Acton and Stlch, 1983) however repair was much less
efficient than for mammalian cell lines. Genotoxlcant Induced
teratogenesls has been observed (Hose et al., 1983; Hose and Puffer,
1983; Hose et_ al . , 1982; Hannah et al . , 1982; Meyer and Jorgenson, 1983;
Muslna and Tsoy, 1981; Rudenko and Tsoy, 1980) though few studies have
related genetic damage to pyknosls and subsequent teratogenesls. An
excellent link has been shown between chromosome damage (metaphase) and
frog embryo abnormalities (McKlnnel, Plcclano and Schaad, 1979) and

- 297 -

breaks In teratogenic embryonic mouse limbs (Loch-Caruso and Baxter,
1984).

Briefly, cell death appears to have been at least partly due to
chromosome damage, In particular multiple damage. Though extensive,
this damage caused few mortalities until 9 days post-treatment, at which
time delayed mortality occurred in the higher concentration groups.
This delayed mortality may be related to cell death, while the induction
of teratogenesis almost certainly was related to observed cytotoxicity.

- 298 -

GENERAL DISCUSSION

The high mitotic rate of the embryo makes the use of anaphase and
diicronuclei analysis widely applicable. Genetic damage in embryos from
polluted water bodies or exposed in Che laboratory may reflect either
parental or embryonic exposure. The sensitivity of the embryo may
reflect its mitotic rate, making Its' use more desirable than the
limited number of adult fish having karyotypes acceptable for metaphase
analysis. This approach provides a technique for assessing genetic
damage which Is broadly applicable to all species of fish.

The passage of contaminants to offspring via yolk Is well
documented (Hose et al_- ; 1981, 1983; van Westernhagen, Rosenthal,
Dethlefsen, Ernst, Harms and Hansen, 1981; Westln, Olney and Rogers,
1983) and constitutes exposure via the yolk. Genotoxlclty evident In
embryos may also be inherited, due to alterations In the parental
genome, for example dominant lethality (if the damage is severe enough),
or genotoxlclty and teratogenicity (Hose _et_ al . , 1981; Hose and Puffer,
1983; Musina and Tsoy, 1981; Rudenko and Tsoy, 1980). Genotoxlc effects
in embryos due to water-borne contaminants have been shown by this and
other studies (see Hose et^ £l . , 1983; Longwell and Hughes, 1980). The
relative importance of these exposure routes is largely unknown, however
BaP caused genotoxlc damage In embryos exposed to 0.5 ng/ml (ppb), while
a parental exposure of 20 mg/kg (ppm) was necessary to cause similar
effects through both genomic and yolk exposure (Hose et al . , 1983; Hose
and Puffer, 1983).

Yolk and water-borne exposure of embryos would be further
influenced by enzyme alterations and chemical distribution In the egg.
Most lipophilic chemicals would accumulate preferentially in yolk,
hydrophllic chemicals possibly throughout the embryo. Little
information is available on the distribution of mutagens or carcinogens,
although BaP (lipophilic) has been shown to accumulate primarily in
yolk, and to a lesser extent in embryonic tissues (Hose et_ £l • , 1982;
Hannah £t al . , 1982) at a time equivalent to 24 hours In this study.
Embryonic enzymes may further alter levels and distribution, generally
mixed function oxidases (MFOs) producing polar metabolites, which are
the active intermediates for some mutagens, and are detoxified products
for others (Franklin, Elcorabe, Vodicnik and Lech, 1980; Parke, 1981).
Fish embryos produce these enzymes (Stegman and Binder, 1978; 1980),
higher levels being inducible by embryonic or parental exposure to
Inducing agents (Hendricks et^ £l • » 1980). The induction of tumors and
genotoxlclty in fish exposed as embryos due to pro-carcinogens
(Aflatoxin, BaP) provides further evidence that the necessary enzymes
are present (Hendricks, 1982; Wales, 1978; Hose e£ al . , 1983) and
binding of the active intermediates to DNA has been shown (Croy, Nixon,
Sinnhuber and Wogan, 1980). Enzyme level fluctuations may in part be
responsible for the variable age sensitivity noted in this study. While
the mechanism for the detoxification of EMS is unknown, MFO levels have
been shown to rise when the liver is formed, though extra-hepatic
tissues also have considerable levels (Binder and Stegman, 1978; 1980).
Increased levels of detoxifying enzymes in embryos exposed initially as

- 299 -

blastulas may have been responsible for the lower levels of genotoxic
damage .

The sensitivity of the genotoxicity endpoints monitored in this
study surpassed that of published studies with adult fish, possibly
reflecting the mitotic rate and/or Insufficient repair, due to the short
cell cycle. Further study of the origin of the genotoxic lesions seen
in this study is warranted. Chromosome fragments observed at metaphase
are due to breakage, however both breakage and "sticky" effects are
reflected at anaphase. The relationships between various lesions at
anaphase, their sub-chromatid events and post-mltotlc results (i.e.
aneuploidy) need to be investigated. The fish embryo may provide a
model for this type of work as up to 75% of the anaphases observed
were damaged, and one possible approach could Include a hanging drop
preparation utilizing yolk-sac cells and phase-contrast optics, which
may allow visualization of individual cells passing through metaphase
and anaphase (Gaulden, 1982; Longwell, 1978). Further, the origin of
raicronuclel must be better characterized, their induction by cyto-
toxicity (due to MIK) raising doubts about their origin due to acentric
fragments or lagging chromosomes, which were not induced by MIK.
Reports of erythrocytic mlcronuclei in juvenile fish (Dr. A.C.
Longwell, personal communication) exposed to mutagens In the laboratory
or in the field must be viewed In light of this apparent anomaly.

The lack of differentiated tissues in the early embryos may be a
disadvantage when assessing tissue-specific carcinogens. The presence
of MFOs, which activate/detoxify these chemicals may be insufficient to
detect chemicals requiring very specific sites and/or other enzymes for
their activation to reactive metabolites.

The ultimate effects of genotoxic damage were varied, including
cytotoxicity mediated teratogenicity, and presumably heritable mutations
(epidermis). It is apparent that most of the observable lesions (AAs)
lead to cell death, as evidenced by the large numbers of pyknotic/
karyorrhexic cells. It is however not apparent whether cell death is
due to division arrest or genotoxicity induced apoptosis. The relation-
ship between embryo genotoxicity and carcinogenicity requires further
study. This work obviously resulted in potentially large numbers of
mutated cells (AAs and epidermal mutagenicity), but no tumors were
observed. This may in part have been due to the delayed mortality In
high concentration groups. The embryo system may prove valuable In
carcinogenicity trials In that the observations of various types of
chromosome damage can be assessed as to their relationship to carcino-
genicity. While genotoxicity has been detected In fish from the Rhine
R. (Allnk et^ al., 1980; Preln et^ al . , 1978; Hooftman and Vink, 1981) and
Duwamlsh R. (Stroraberg, 1981), only in the latter area have elevated
tumour rates been noted (Pierce, McCain and Wellings, 1978; Mallns,
McCain, Brown, Sparks, Hodgkins and Chan, 1982; McCain et_ al_. , 1982;
Sloof, 1983; Poels, van der Gaag and van de Kerkoff, 1980; Kurelec et_
al., 1981). Studies of embryos from these areas may aid in determining
whether chemical genotoxicity or viruses are responsible for the
observed tumours .

- 300 -

A second potential use is to assess the genotoxic impact of
chemical and physical agents used to induce heritable mutants. Mutants
induced by X-rays in germ cells and pre-gonial (blastula) cells
(Chakrabarti, Streisinger, Singer and Walker, 1983; Walker and
Streislnger, 1983) could be examined for structural defects utilizing
the techniques used in this study-

This study has demonstrated many effects of the alkylating agent
EMS» including induced chromosomal damage in embryonic tissues, and
heritable defects in juvenile epidermal tissues. The embryo approach
has tremendous potential for both laboratory and environmental studies
of genotoxic chemicals, because it can be applied to any fish species.
This permits the evaluation of DNA damage in relationship to other end-
points, most notably cancer. Most studies of aquatic genotoxins to date
have concentrated on in vitro work with bacteria and cell cultures
coupled with the chemical concentration of water samples, or on DNA
damage In only certain species.

Exciting potential applications of the prsent work include its use
in identifying whether a chemical component (reflected as mutations) is
present in environmentally induced tumors, or in the carcinogenicity of
chemicals when administered to embryos. The manipulation of a variety
of modifying factors. Including route of embryo exposure, and modifying
influences of other chemicals make this system potentially adaptable to
a variety of areas of concern, with regard to genotoxiclty . The
observation of genotoxiclty related effects, including teratogenesis,
carcinogenesis, delayed toxicity and inherited defects may lead to a
better understanding of the potential effects of the dilute contaminants
which are biologically integrated and accumulated by fish inhabiting
polluted waters, or acting as a model in laboratory studies.

- 301 -

ACKNOWLEDGEMENTS

The authors wish to gratefully acknowledge the financial support
provided by the Ontario Ministry of the Environment^ from the Provincial
Lottery Fund. We are also indebted to J. Middlemas, S- Brown, Drs. H.
Ferguson and J.B. Sprague, and D. Walker for their help. A special
thanks must be extended to Dr. D. A. Rokosh and Mr. G.R. Craig, liason
officers for this project.

REFERENCES

To avoid an overtly long paper, references have been omitted from
this section, however citations can be obtained from either of the
following articles, or a full listing can be obtained from the senior

author .

Smith, I.R. and V.E. Valli, 1983. The development of a freshwater fish
test to identify aquatic toxic contaminants. Conference
Proceedings, Ministry of the Environment Technology Transfer
Conference #4, Nov. 29,30, 1983. p 409-437.

Smith, I.R., 1984. The development of an aquatic genotoxicity test with
Brachydanio rerio embryos. M.Sc. Thesis, University of Guelph,
Guelph, Ont. pp 116.

- 303 -

FIELD ^EASURE^Em• OF INFILTRATION
THROUGH LANDFILL COVERS

PHASE I

BY:

A.G. Hims: Gartner Lee Associates Limited

P.K. Lee: Gartner Lee Associates Limited

R,W. Gillham: University of Waterloo, Dept. of
Earth Sciences

Paper suixnitted for Presentation at The Fifth Technology
Transfer Conference Sponsored by The Ontario Ministry
of The Environment, November 1984.

OCTOBER, 1984

- 304 -

FitiLD MEASUREMENT OK INFILTRATION THROUGH LANDFILL COVERS

A.G. Hims, - Gartner Lee Associates Limited

P.K. Lee - Gartner Lee Associates Limited

R.W. Gilham - University of Waterloo,

Earth Science Department

ABSTRACT:

This paper reports the final results of the Phase 1 Study
which was to design , construct and field test
instrumentation capable of measuring the infiltration which
occurs through the final cover material at a landfill site.
The study was initiated in 1982 and collection of data has
been on-going since March 1983. Details of the lysi meter
design and construction, together with the initial results,
were reported in the proceedings of the Fourth Annual
Technology Transfer Conference in November 1983- An
analysis of all data obtained between March 1983 and July
1984 is now presented, with an assessment of the overall
outcome of this phase of the study.

- 305 -

FIELD MEASUREMENT OF INFILTRATION THROUGH LANDFILL COVERS

BIT

A.G. Hims
P.K. Lee
R.W. Gillham

Gartner Lee Associates Limited

University of Waterloo
Earth Science Department

ABSTRACT

This paper reports the final results of the Phase I Study which was to design,
construct and field test instrumentation capeible of measuring the infiltration
which occurs through the final cover material at a landfill site. The study
was initiated in 1982 and collection of data has been on-going since March 1983
Details of the lysimeter design and construction, together with the initial
results, were reported in the proceedings of the Fourth Annual Technology
Transfer Conference in November 1983. An analysis of all data obtained
between March 1983 and July 1984 is now presented, with an assessment of the
overall outcome of this phase of the study.

INTRODUCTION

One of the objectives of the final cover which is applied to a completed
Icindfill is to minimize the amount of infiltration which occurs through the
cover into the landfill. The eunount of leachate which is generated within
most landfills is generally equal to the amount of infiltration which occurs,
once the wastes have reached field capacity. By minimizing the amount of
infiltration through the cover material, the rate of leachate generation is
also minimized. This is desircible fzom both an environmental and leachate
treatment cost point of view.

Present cover design practice utilizes both empirical and agricultural
drainage equations together with climatic factors in order to estimate

- 306 -

infiltration and thus predict leachate generation rates. The usual method
of estimating infiltration is to use a water budget approach by applying the
following equation:

Infiltration = Precipitation - Runoff - Evapotranspiration ±
Change in Soil Moisture Storage

Although the equation appears simple, the determination of each parameter is
by no means straightforward and can give rise to significant errors.

In Canada in particular, there are no direct field measurements of the long-term
infiltration characteristics through covers on existing landfills that would
demonstrate that these designs are valid and that infiltration is as low as
predicted by the water budget equation. This is an important issue, not only
when the siting of a new facility is considered, but also when the impacts of
existing or closed-out landfill sites are being assessed. The rate of leachate
generation is obviously a key factor in the design of on-site collection and
treatment facilities for leachate. Also, the prediction of the off-site intact
of a contamincint plume, including the costing and design of contingency plans,
as required by the Ministry of the Environment (MOE) is dependent upon a
knowledge of the leachate generation rate.

In light of the foregoing, the Minist^ of the Environment is funding a two
phased study to measure the amount of infiltration which occurs through the
final cover material at landfill sites located in various physical settings.
This paper reports the results arising from Phase I of the study.

PURPOSE AND SCOPE OF PHASE I

The main objective of Phase I is to design, construct and field test an econo-
mical and practical field lysimeter installation that will provide accurate
and reproduceable measurements of infiltration through a final cover material
at an existing landfill site.

- 307 -

More specifically, the purpose and scope of Phase I is as follows:

a) to provide a preliminary description of the special field
tests and experiments to be undertaken in Phase 2,

b) to select and obtain access to a site to be used in Phase I,

c) to design, select and specify equipment.

d) after approval of the design and location of the test
site.- given in writing by the Crown, to construct the
test infiltration collection systems for performance
testing,

e) to bring the infiltration collection systems to field

capacity or into a condition such that useful measurements
can be obtained.

fj to carry out performance testing.

If) to carry out a performance assessment of the test
infiltration collection systems.

SITE LOCATION

The site chosen for the installation of the lysimeters was the Britannia
Road Landfill Site, Mississauga, located in the Region of Peel.

- 308 -

This site is an engineered facility with cells excavated and bottomed in low
permeability Halton silty-clay till subsoils. The site layout and location
of the lysimeter installations is shown on Figure 1.

B,,linn>*

■»0' '0 SCitt

BRITANNIA ROAD LANDFILL
CELL LAYOUT

F.g

MEASUREMENT OF INFILTRATION
Lcgtnd

Croiiti 8? i'i

This landfill is constructed as a leachate containment site with provisions
for leachate collection by means of a network of underdrains within the
base of individual cells. The underdrains feed into a perimeter collection
system which removes the leachate off-site to a sewage treatment plant.
In the current method of operation, the leachate which forms within each cell
is continually removed from that cell by gravity flow to the underdrains.
Leachate is not allowed to mound up into the refuse to any great degree. .

- 309 -

The total volume of leachate which passes through the perimeter system
to the treatment plant is monitored. Monthly values of leachate volumes
were provided to Gartner Lee Associates Limited by Peel Regional Staff.
It should be noted that the sewage frcnn the scale house and site building
also contributes to the total volume of leachate passing through the
sewer. At the outset of the study. Cells 1 and 2 were ccmpleted eind
had the final cover of reccmpacted clay till in place. Cell 1 was also
topsoiled and vegetated. Cell 2 was topsoiled, and was seeded in
September 1983, during the study period. Landfilling was taking place
in Cell 3 initially and Cell 4 came on-stream in March 1983. The exca-
vation and underdrain system for Cell 4 had been mostly completed prior
to initiation of the study. As a consequence, any precipitation which
fell into the cell would enter the leachate collection system and would
thus contribute to the volume of leachate which was measured at the
monitoring station.

Each daily lift of garbage is covered with soil to minimize litter and
rodent problems. When the cell has reached the final height, the
final cover material is placed in accordance with the overall final
grading plan. The final cover consists of the silty clay till soils
which has been excavated from the next cell to be landfilled. The soil
cover is spread by scrapers and final graded by means of a bulldozer.

LYSIMETER DESIGN AND CONSTRUCTION

Details of the lysimeter design and construction techniques which
were used are reported in Hims et al 1983 and 1984. Two designs
of lysimeter were constructed at the site. However, results have only
been obtained frcxn lysimeter design #1, general details of which are
presented in Figure 2 .

- 310 -

rtiMt

SCALE 1 85

The design of each lysimeter basically consists of an infiltration capture
unit from which the water is directed to a storage well equipped with an
automatic water level recorder. The original design called for the lysimeter
to be constructed below the cover material, within the garbage. However, at
this particular site, the thickness of the final clay cover (2-3m) was signi-
ficantly greater than anticipated. As a consequence, the lysimeters were
constructed within the cover material, with only the storage well penetrating
into the underlying garbage.

The storage well was installed by means of a backhoe because it was not
possible to penetrate the garbage using a Stirling auger drill rig, as was
originally planned. The size of the excavation required to position the
storage well did result in some differential settlement problems following
construction. The connection between the storage well and the lysimeter
severed on two lysimeters and required reconstruction.

- 311 -

The surface area of the lysimeter was made lOm* in order to be repre-
sentative of the general soil cover conditions. It should be remembered
that as the garbage decomposes, differential settlement will occur within
the cover soil. This gives rise to fracturing within the cover which
creates a secondary permeability within the soil. This fracture permeability
can ultimately become the major factor which controls the amount of infiltration
occurring through the cover.

The volume of storage available in the well, up to the level of the
discharge pipe from the lysimeter, is equal to 0.47 m' . This is equivalent
to a total infiltration of 47 mm over the surface of th^ lysimeter
installation. The sensitivity of the installation is such that an infil-
tration of 1 mm into the lysimeter should produce a water level rise of
57.6 mm in the storage well, assuming 100% efficiency of the instrument.
The automatic water level recorders have a working range of approximately
3 m. The minimum water level increase in the well which can be identified
is about 5 mm. Thus the minimum amount of infiltration which could
realistically be observed by this design is theoretically 0.09 mm.
However, in practice, the response of the lysimeters was found to be less
sensitive than this.

Since it had been necessary to pump out water from the collection wells on
several occasions, both during and after construction, it was assumed that
the installations could be considered to be at field capacity by the time
the recorders were operative. Removal of water from the storage wells is
accomplished by means of a contractors sump pump, equipped with a 6 m intake
line. The ground surface at the installations remained without topsoil until
the area was seeded in early September.

- 312 -
The actual layout of the five lysimeters which were constructed is illustrated
on Figure 3.

IJJm

■6'/.

4
>
1

i» Zm

t\Jm

Ottif" I

- »

inic

ItdtCV p4p»

•'™ • iMiatt ■•"

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12>M • px

• JJ"

'SI.

• i2 m

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M5

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CtLk. f

Figure 3

Field Lay-Oul Of Lvsimeter Installation

INFILTRATION RESULTS:

DATA COLLECTION

Data regarding infiltration into the lysimeter installations have been
generated from January 1983 onwards. Continuous water level measurements in
the three storage wells have been obtained from mid-March 1983 and collection
of data are on-going. Daily summary records of climate data have been obtained
on a monthly basis from the meteorological station located at Lester B. Pearson
International Airport, (Toronto) , which is situated approximately 12 km to the
northeast. Records of the total volume of liquid {leachate plus waste water
sewage) which is pumped from the site to the sewage treatment pleuit have been
provided by the Region of Peel on a monthly basis.

- 313 -

Water Level Response Curves

An example of the water level response curves, from the three lysimeters is

presented on Figure 4.

(■iuurr 4

I \si\U riK

Witcr ttiei Rtipimtc In Slorigf Wells

l.\M.MI II H 1

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imt

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The daily total precipitation data are also provided on the charts. The raw

data charts cind the precipitation records are not presented in this paper,

but are kept on file with the Ministry. The shape of the response curves

presented on Figure 4 is generally typical of that obtained for the majority

of the study period, with the exception of the summer and early fall.

The lysimeters respond to a rainfall event relatively quickly, depending on

the size of the event and antecedent rainfall. The initial response is a

rapid increase in water level in the storage well followed by a much slower

increase after the event has finished. The tapering-off effect is likely

due to gravitational drainage out of the lysimeter backfill, once the infiltration

process has stopped.

- :^i4 -

The responses obtained during the late summer - early fall period have been
much sharper or stepped in appearance, in as much as the rounded, tapering
off part of the curve is not present. Each sharp rise in the water level is
followed by a flat section indicating no further increase in the level until
the next rainfall event of any significance. Also, the magnitude of the
response is not as great during the summer period. This is probably due to the
drying out of the soil cover and lysimeter backfill material in between rainfall
events and the retention of the initial portion of infiltration within the backfill
as the moisture content returns to field capacity.

A problem arises in analyzing the response curves if the water level in the
storage well is allowed to rise to the level of the discharge outlet from
the lysimeter into the well. This has occurred on five occasions during the
study in April, November-December, 1983, and February, April and May 1984.
The response curves indicate that the water level in the storage wells
apparently fluctuates up and down, sometimes by as much as 0.3-0.5 m or more.
The cause of the fluctuations may be a combination between additional infiltra-
tion during rainfall events, and possible leakage at the outlet pipe.
Response to barometric changes may also create minor fluctuations.

LYSIMETER CALIBRATION

The calibration exercise was carried out in the late spring of 1984 at which
time the lysimeters had been in operation for approximately one year.
The calibration was carried out by applying a known volume of water to the
surface of each of the three lysimeters on a number of occasions and recording
the water level response in the storage wells. The water was applied slowly
from a 45 gallon drum (204 L) by gravity flow from a garden sprinkler hose snaked
across the ground surface. The water came out of the hose in droplet form,
rather than as a stream. Plastic garden edging was placed around the perimeter
of each lysimeter to prevent surface runoff from both entering the test area
from upslope, and from actually running off the test areas. Canopies were
constructed from plastic sheeting over each test area with the intent of
preventing interference from rainfall events during the test. The canopies
were only partially successful.

- 315 -

Figure 5 presents the graphical plot of the water level response in the three
lysimeter storage wells during the calibration exercise. Table 1 sxmmarizes
the interpretation of the data extracted from these response curves.

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Lysimeter Calibration Results

Table 1

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The response to the initial application of 68 litres was extremely poor.
This has been attributed to the high water level in the storage well at
the outset of the calibration. Throughout the study period, it has been
observed that once the water level in the wells approaches the discharge
outlet pipe from the lysimeter, the lysimeters cease to respond in the
expected manner. The response to rainfall events is greatly subdued,
indicating that the instrument is effectively full or saturated, and also
leakage has previously occurred at the pipe connection to the well from
all three lysimeters .

The interpretation of the results from the calibration exercise is very
subjective, particularly for tests one and two. The only apparently
reasonable results were obtained from the second calibration test.
Based on the assumption that the response is observed in the storage
wells was due to the water applied during the first and second tests,
the response appears to range between 70 to 92% of the volume applied.

The first test probably did not produce the expected response because the
storage wells were full of water prior to the test. The cause (s) of the
very low responses observed following test 3 are not certain. One problem
that was common to all tests was the very slow application rate achieved by
the sprinkler hose method. This could result in significant losses due to
evaporation caused by wind blowing beneath the protective canopies. Also,
since the grass is well established on the surface of the lysimeters, plant
uptake could also remove a portion of the water.

In sxmimary, the results of the calibration tests do not provide conclusive
evidence to indicate whether or not the lysimeters are functioning as designed,
This aspect is a very important part of the project, and the calibration
should be attempted again. Until the calibration is completed, the results
obtained during this phase of the project should be treated with caution.

- 317 -

COMPARISON OF RESULTS FROM THREE LYSIMETERS DURING THE STUDY

Since the calibration exercise was not conclusive, the field response charts

have been analyzed at face value.

TaJDle 2 summarizes the infiltration data on a monthly basis extracted from

the field response charts. Table 2 also presents the monthly volume of leachate

which was pumped from the leachate collection system during the study period.

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Figure 6 presents these data graphically:

VakjM Of tytknatar fciflltntlon. Monthly Prvclpilatlon And Total l^eaclut* Volumes For
Britannia Road Landfitr Site , —

• ••■-•4

MM CiH

..I"

FIQURE 6

- 318 -

Four levels of comparison are available to assess the relative consistency
of the results from the three lysimeters. These are: (a) on a rainfall event
by event basis, (b) on a monthly basis, (c) comparison over the entire study
period, and (d) a comparison with the leachate volume pumped through the
collection system.

COMPARISON ON AN EVENT BASIS :

This was accomplished by assessing the magnitude and shape of the water level
response curves arising from individual rainfall events for the three lysimeters,
However, since the individual monthly data charts are not appended here, only a
summary of the comparison will be presented.

In summary, it is very difficult to assess the performance of the lysimeters
on an event by event basis. Each lysimeter appears to respond somewhat
differently during the various seasons on the year. Lysimeter 1 generally
produces a smooth well rounded response curve which does not reflect individual
events. Lysimeter 2 has a more stepped response curve which does show
individual events in some cases. Lysimeter 3 lies somewhere in between.
Another problem which has come to light in assessing individual responses
is the time lag effect, both before a response is observed and following the
rainfall event. During parts of the year the responses are very quick,
within hours, but at other times the response is delayed by up to one day or
greater. Also the response following an event may not be completed before
another event occurred which masked the previous response.

COMPARISON ON A MONTHLY BASIS :

Table 3 presents the infiltration recorded in the three lysimeters on a monthly
basis throughout the study period, from April 1983 to July 1984. The infil-
tration data are presented in terms of the percentage of the total precipitation
which occurred during the month. The mean value of the results from the three
lysimeters is presented together with the standard deviation and coefficient
of variation.

- 319 -

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Figure 7 illustrates graphically the infiltration measured in the three lysimeters
and that estimated from the leachate volume passing through the collection system,
all expressed as a percentage of the total precipitation.

Infiltration In Three Lysimeters And Leachate Collection System
Expressed As % Of Total Precipitation

Figure 7

- 320 -

The standard deviation and coefficeint of variation show a wide range in

value during the study, and the correlation between the results from the three

lysimeters does not appear that good. For example, the coefficient of variation

shows a range from a low 3% in February 1984, which indicates a very good

correlation, to a high of 112% in July 1983, which indicates an extremely

poor correlation. Also, the coefficient of variation is less than 25%

for five out of sixteen months, and it is greater than 75% for three months.

However, the data indicate that a reasonable degree of correlation exists
when the infiltration into the lysimeters exceeds 30% of the total monthly
precipitation. For the seven months when this occurs, the coefficient of
variation is below 30% for six of these months, and the average value for the
seven months is 19%. A much greater variation in the results exists when
the infiltration is below 30% of the total monthly precipitation, which
occurred in the summer of 1983 and 1984, and January 1984 when snow cover was
present. Thus, even though these results show a wide variation, relative
to one another, in terms of the total amount of infiltration recorded over
the sixteen month period, they do not have a pronounced effect, becuase the
total volume of infiltration during this period is not large (61 mm out of a
total of 353 mm during the study period) .

COMPARISON OVER TOTAL STUDY PERIOD:

The total cimount of infiltration recorded in the three lysimeters during the
period April 1983 to July 1984 are presented on Table 3. The values, expressed
in mm of infiltration are as follows:

LYSIMITER

LYSIMETER

MEAN

STANDARD

COEFFICIENT
OP

Total

VALUE

DEVIATION

VARIATION

Infiltration

(mm)

382.7

345.0

332.2

353.3

26.3

7.4

Total

Infiltration

as % of Total
Precipitation

35.9

32.4

31.2

33.2

2.4

7.4

- 321 -

These data indicate the correlation between the three lysimeters, in terms
of the total eunount of infiltration which occurred during the study period, is
good. This conclusion really follows on from the previous section which indicated
that a reasonable correlation exists between the three lysimeters when the
monthly infiltration exceeds 30% of the monthly precipitation {spring and fall) .
Since this infiltration accounts for approximately 83% of the total infiltration,
the above result appears reasonable.

The mean value of total infiltration which occurred during the study
period is 353.3 mm. This is equivalent to 33% of the total precipitation
which occurred during this period.

COMPARISON WITH LEACHATE VOLUME PUMPED THROUGH COLLECTION
SYSTEM :

Table 2 presents the monthly volume of leachate pumped through the collection
system.

Interpretation of these data is not straightforward, particularly when attempting
to relate these figures back to an amount of infiltration through the landfill
cover. A problem arises due to the fact that only cells 1 and 2 are completed
and closed, and that landfilling has been active within cells 3 and 4 during
the study period. During the initial stages of the study, landfilling was
carried out in cell 3 and the majority of cell 4 was excavated, with the under-
drain system in place. Landfilling was commenced in cell 4 in mid-March 1983.
Thus a large portion of the precipitation falling onto cell 4 during the period
January to March, would enter the underdrain collection system and contribute
to the total volume of leachate recorded at the pump station. The following
assumptions were adopted into the analysis of these data:

- 322 -

1) The area of cell 4 is 6.5 ha, the area of cells 1 and 2 is
15 ha total.

2) 100% of the precipitation falling onto the open cell 4 excavation,
during the period January - March 1983, enters the xinderdrain
system.

3) 100% of the precipitation falling onto cells 3 and 4 while landfilling
is active is either evaporated back to the atmosphere or is absorbed
by the refuse.

4), The refuse within cells 1 and 2 was at field capacity prior to the
study .

5) The underdrain collection system is 100% efficient in removing
all leachate from the cells.

6) The sewage contribution to the total leachate volume is 60 m*
per month.

7j There is no ground water inflow into the cells.

Based on the above assumptions, and the volume of leachate pumped through the
collection system, the infiltration which occurred through the landfill covers at
cells 1 and 2 has been estimated. The results are presented on Table 2 and also
Table 3, expressed as a percentage of the total monthly precipitation.
These data are presented graphically on Figure 7.

Figure 7 shows that whereas the estimated infiltration based on the volume
of leachate pumped through the collection systems follows a similar trend to
the infiltration recorded in the three lysimeters, there is an apparent time
lag in the reponse of approximately four months. If the plot for the leachate

- 323 -

collection system is moved back four months, the relative curve match
becomes good. Table 3 present the data, with the four month time lag
taken into account.

In terms of the total volume of infiltration, which occurred during the
study period, based on the volume of leachate, both methods of cinalysis yield
the same result. These results indicate that approximately 25% of the total
amount of precipitation which fell onto the landfill site infiltrated through
the final cover material and into the refuse to become leachate. This figure
compares very favourably with the mean value of 33% obtained from the lysimeters
The four month time lag in response may be associated with the travel time
of the leachate from the landfill surface to the collection system.

DISCUSSION

The three lysimeter installations have been monitored over a sixteen month

period. An attempt to calibrate the lysimeters in April and May 1984 was

unsuccessful due to a variety of factors. Only one part of the calibration

yielded encouraging results, when the efficiency of the lysimeters was

estimated as ranging between 70 to 92%. Additional testing under more

suitable conditions may yield results which are more conclusive.

The individual lysimeters do not show much of a correlation on an event

by event basis. Each lysimeter responds in a different fashion both frcan

one another and seasonally. Lysimeter 1 generally shows a fairly lengthy

tapering off period following an infiltration event. Lysimeter 2 usually

shows a more stepped response curve with very little tapering off and Lysimeter

3 lies somewhere in between. This difference in individual response patterns

may be explained in terms of the variation in lysimeter backfill and soil

cover material. Even though every effort was made to construct the lysimeters

in the same manner, some degree of variation should be expected. Factors

which may affect the response pattern include; (a) the natural variation

in the refuse which was used as backfill in the lysimeters, (b) the variation

in the material and the compaction achieved in the soil cover, and, (c)

- 324 -

variation in the affect of the grass vegetation on the surface water run-off/
retention characteristics.

The correlation between the infiltration recorded in the three lysimeters
on a monthly basis is relatively good, particularly when the infiltration is
greater than 30% of the monthly precipitation. The range in the coefficient
of variation during these periods is between 3 and 38%, and the mean value
is 19%. Also, approximately 83% of the total infiltration which occurred
during the study was recorded during those months when infiltration exceeded
30% of the total precipitation.

A much greater variance is apparent in the results from the three lysimeters
when the infiltration is less than 30% of the total precipitation. The reason
for this variation, which occurs during the summer months, is probably
associated with the drying out of the soil cover and lysimeter backfill ernd
subsequent re-wetting during a precipitation event. Since only about 17%
of the total infiltration which was recorded during the study period occurred
during those months with the greatest variance in results, the effect of that
variance is not significant.

The correlation between the results from the lysimeters for the entire

study period is very good. The mean value of the total infiltration recorded

during the sixteen month period is 353.3 mm, the standard deviation for the

three lysimeters is 26.3 mm and the coefficient of variation is 7.4%.

In terms of the total amount of precipitation which fell during the same period

the infiltration is equivalent to 33%. This figure correlates very well with

the value of 25% which is the estimate of infiltration based on the total volume

of leachate {exclusive of sewage) which was pumped through the collection

system during the study period. Again, it should be emphasized that these

results are very preliminary in nature, and are not intended for use in

landfill design.

- 325 -

SUMMARY

The results obtained indicate that the three lysimeter installations
have for the most part met the overall objective of the study. That
objective was to design, construct and field test an economical and
practical field lysimeter installation that will provide accurate and
reproduceable measurements of infiltration through a final cover material
at an existing site. The large variation which was encountered in the
results during the summer period is not unexpected, and does not have a
pronounced effect on the total volume of infiltration which was recorded.

The leakage which seems to occur when the water level in the storage well
reaches the elevation of the discharge pipe from the lysimeter is a concern.
This may indicate that the pipe is partly severed at the coupling which,
in time, will cause the instrument to malfunction completely. We suggest
that the coupling be uncovered by careful excavation and repaired. Also,
since the design of the lysimeter s appears to function satisfactorily, it
may be advantageous to carefully re-excavate at least one of the installations,
In this laannez, it will be possible to inspect the key parts of the lysimeter
cind assess their performemce following almost two years burial. In-situ
density testing should also be carried out at various depths in the soil cover
in order to assess the degree of compaction which has been achieved.
Following this inspection of the lysimeter (s) , the installations should be
reinstated and monitoring continued. Also, the calibration of the lysimeters
should be attempted again under more suitable field conditions.

ACKNOWLEDGEMENTS :

We wish to acknowledge the guidance provided by Dr. G. Hughes who acted
as Liaison Officer for the Ministry of the Environment. We also wish to
acknowledge the Region of Peel for allowing us to construct the lysimeters
at the Britannia Road Landfill and to thank Mr. L.G. Conrad, Department of
Engineering, euid Mr. M. Walters, Site Supervisor, for their assistance during
the construction eind monitoring portions of the program.

REFERENCES

Fenn, D., Hamley, K., DeGeare,T. 1975. Use of the Water Balance Method for
Predicting Leachate Generation from Solid Waste Disposal Sites. (U.S.
Agency 530/SW-168) .

Fungaroli, A., Steiner, R.Lee. Investigation of Sanitary Landfill Behavior.
Volume 1 Final Report, Voliame 2 Supplement to Final Report. Drexel
University, Philadelphia, Penn. 19104. Municipal Environmental Research
Laboratory Office of Research and develofMnent U.S. E.P. Agency.
Cincinnati, Ohio, 45268. Available through NTIS.

Gee, J.R. 1983. The Prediction of Leachate Generation in Lemdfills. A New

Method. Sixth Annual Madison Conference of Applied Research and Practice
on Municipal and Industrial Waste. Dept. of Engineering & Applied Science.
University of Wisconsin - Extension. September 1983.

Gee, J.R. 1981. Prediction of Leachate Accumulation in Sanitary Landfills.

Fourth Annual Madison Conference Applied Research and Practice on Munici-
pal and Industrial Waste. Dept. of Engineering and Applied Science,
University of Wisconsin - Extension. 432 N.Lake Street, Madison, WI 53706.

Hims, A.G., Lee, P.K. and Gillham, R.W. 1984 Field Measurement of Infiltration

Through Landfill Covers. Presented at the Seminar on Design and Construction
of Municipal and Industrial Waste Disposal Facilities. Sponsored by the
Canadian Geotechnical Society and the Consulting Engineers of Ontario.
June 1984.

Hims, A.G., Lee, P.K., Gillham, R.W. 1983 Field Measurement of Infiltration

Through Landfill Covers. Fourth Technology Transfer Conference sponsored
by Ontario Ministry of the Environment, Toronto, Ontario. November 19B3.

Lewis, M.R., Powers, W.L. 1938. A Study of Factors Affecting Infiltration.
Soil Science Society of American Proceedings 1938.

Lutton, R.J., Regan, G.L. & Jones, L.W. 1979. Design and Construction of Covers
for Solid Waste Landfills. Army Engineer Waterways Experiment Station;
Vicksburg, Mississippi. U.S. EPA/600/2-79-165.

Mather, J., Rodriquez, P. 1978. Use of Water Budget in Evaluating Leaching
Through Solid Waste Landfills. Delaware University, Newark. Office
Worker Research and Technology (NITS) .

Proctor S Redfern Limited. 1977. Design Report for Central Britannia Road
Landfill Site (Site 4) for the Regional Municipality of Peel.

- 327 -

Cuinlan, P., Bunnem, R., Siemer, E. 1982. In-Situ Lysimeter Installation,
Presentation at Summer Meeting Americeui Society of Agricultural
Engineers. University of Wisconsin-Madison.

Stegman, R. 1979. Leachate Treatment at the Sanitary Landfill of Lignen,
West Germany: Experiences with the Design and Operation of the
Aerated Lagoons. Second Annual Madison Conference Applied Research
and Practice on Municipal «ind Industrial Waste. Dept. of Engineering
and Applied Science, University of Wisconsin-Extension. 432 N, Lake
Street, Madison WI 53706.

- 329 -

Development of Specific Protein Adsorbents for Selective Extraction
of Trace Contaminants Significant to Human Health:
Modelling of Fetal Cross-Placental Uptake Specificity

Carleton J. C. Hsia

Department of Biochemistry, Faculty of Medicine
University of Toronto, Toronto M5S 1A8

- 330 -

Development of Specific Protein Adsorbents for Selective Extraction

of Trace Contaminants Significant to Human Fetal Health:

Modelling of Fetal Cross-Placental Uptake Specificity

Department of Biochemistry » Faculty of Medicine,
University of Toronto, Toronto M5S 1A8

The complementary ligand binding specificities of maternal serum albu-
min and fetal alpha-f etoprotein (AFP) for nutrients e.g. polyunsa-
turated fatty acids (PUFA), and metabolic wastes e.g. bilirubin, have
been proposed to provide a specific mechanism for the transport of
these ligands across the maoMnalian placenta. We have developed an in
vitro assay to demonstrate the principle of this mechanism. This
assay shows that in the presence of AFP and albumin, PUFA as well as a
known cross-placental teratogen, dlethylstilbestrol (DES), bind
specifically to AFP. Using radiolabelled fatty acids and DES, we have
shown that 1 umole of AFP in a ConA-Sepharose column retains greater
than 95'/. of PUFA and 90^ of DES when nanomolar solutions of these
ligands pass through the column. We propose that AFP is responsible
for the uptake and concentration of both PUFA and DES in the fetal
circulation. On the assumption that other biohazardous substances
which have high AFP-binding affinity are likely to be similarly con-
centrated in the fetus, we believe that the binding specificity of AFP
can be used for the extraction of trace environmental contaminants
significant to human fetal health.

- 331 -

INTRODUCTION - THE MODEL

Thf mamm^il ian fetus is a complex organism whose extremely rapid
growth gives it stringent nutritional requirements. The source of its
nutrition is the maternal blood; the fetus depends for its nutrient
uptake and waste disposal entirely upon exchange between maternal and
fetal blood at the placenta, and the placental membrane which
separates the two circulations provides a large area for this exchange
(I). Although nutrients, wastes, and many drugs are known to cross
the placenta, the mechanisms of such transport are obscure; simple
diffusion of most substances is assumed (2).

An exception to this is the transport of oxygen to the fetus.
A steady suppl y of oxygen is absolutely essential to the fetus and it
has been establ ished that there is a mechanism which allows enhanced
fetal uptake of oxygen over a wide range of oxygen concentrations in
the maternal blood. A fetal variant of hemoglobin (HbF) exists (3)
which, due to its structure (4,5) and the pH of fetal blood (6,7) has
a higher affinity for oxygen than does maternal hemoglobi n (HbA) .
r,i ven that oxygen can diffuse through the placental membrane just as
it does through the alveolar membranes of the lung, this suggests a
mt-chanisin for fetal oxygen uptake beyond simple diffusion: as oxvgen
dif fuses -across the placental membrane a low concentration of t ree
oxygen is maintained on the fetal side, where it is more tightly bound

- 332 -

than on the maternal side. A gradient of free oxygen concentration is
therefore maintained across the placenta, down which oxygen flows from

the maternal to the fetal blood.

The oxygen affinity of fetal blood var ies throughout ges tat ion
with the i:oncent rat ion of fetal hemoglobin. HbF concentration and the
oxygf^n aiiinity of fetal blood are greatest in early gestation, and
de.rease as adult hemoglobin is expressed (8,9). Fetal uptake of oxy-
g«'n by thf-' HbF-HbA lacilitated diffusion is greatest, therefore, dur-
ing early gestation when organogenesis occurs.

A*, analoj^ons mechanism has been suggested more recently for
rross-pla.ental transport of plasma -borne 1 igands , involving serum
.^ ] biimiii ami alpha- fetoprotein (AFP) . Bilirubin, fatty acids (10,11),
steroids ( ! 2 , 1 "S ) , and hydrophobic drugs ( 1 A ) , are bound and tran-
sjKTted in the blood of adults by serum albumin. AFP is a fetal
gl vioproL-" in wh i ch is related to a I bum i n , wi th a common ancestral gene
( ] ') ) and similarities in amino ac:id sequence (16), immunologic reac-
tivity (\7), and phys i cochemica 1 properties (18). It is thought to
liave a ftmrtion in the fetus similar to that of albumin in the adult,
iiamnl y t ' nispor L of 1 igands and maintenance of osmot ic pressure in the
hi ood .

Th.- finding that AFP specificallv binds polyunsaturated fattv

- 333 -

acids (PUFA) much more strongly than albumin suggested that albumin
and AFP art analogously to maternal and fetal hemoglobin to concen-
trate the essential PUFA in the fetal circulation (19). (Similarly,
albumin binds bilirubin more strongly than AFP (20) » suggesting a
mechanism for excretion of this waste by the fetus into the maternal
blood.) A model in which the different 1 igand-binding affinities of
albumin and AFP mediate a facilitated diffusion of PUFA into the fetus
is now widely accepted.

THE HYPOTHESIS

This model provides for fetal uptake of ligands beyond simple
diffusion. Enhanced and selective uptake (and possibly» in the case
of bilirubin, excretion) of ligands seems necessary to satisfy the
special niitr 1 tiona] needs of the fetus , giving it a nutritional advan-
tage throughout a range of nutritional states in the mother . We
hypothesize that this mechanism can act generally on ligands with high
affinity for AFP» including exogenous substances which can also be
concentrated in the fetus by albumln-AFP exchange. According to this
hypothesis, for example, a drug for which AFP has a higher affinity
t.h*iTi albumin might be present in the maternal circulation at a thera-
peutic concentration, but be concentrated to toxic levels in the
letuK . ( Further evidence for AFP-mediated 1 igand uptake has come from
a recent study of estrogen uptake by the rat fetus (25): the fetal
uptake of a several natural and synthetic estrogens injected into the

- 334 -

mother corresponded closely with their AFP-blnding affinity.) Con-
versely, a drug for which AFP has low affinity could be sequestered In
the maternal compartment . It might be this selective uptake mechanism
which potentiates the effects of some teratogens; if this is true,
then knowledge of the relative affinities of albumin and AFP for a
given llgand should allow prediction of its teratogenic potential.

THE ASSAY

To test this hypothesis , we have devised a chromatographic
assay of the relative binding specificities of albumin and AFP which
simulates li gand exchange at the placenta. In this assay, a trace
amount of the radiolabelled 1 igand of Interest is added to a mixture
of al bumln and AFP. The al bum! n-AFP-1 igand mixture is then passed
through a column containing a ConA-Sepharose affinity gel. AFP, with
attauVied 1 i gand, is retained by the column due to its sugar moiety
while albumin and any ligand associated with it are eluted. AFP is
subsequently eluted by a sugar-containing buffer. The radioactivity
associated with the two separated proteins indicates the relative
aftinity ot each for the ligand.

The ConA-Sepharose assay offers a good approximation in vitro
to our view of the mechanism of the transplacental exchange of some
ligands (i.e. those with high affinity for AFP). As the proteins and
\ i gand pass through the column, the 1 igand is in equilibrium between

- 335 -

specific binding to elCes on albumin and AFP and non-specific binding
to the ConA-Sepharose matrix. The matrix Is analogous to the placen-
tal membrane compartment; diffusion into and out of this compartment
is a function of the binding affinities of the proteins and the solu-
bility of the ligand in the various compartments. Separation of the
two proteins In the assay is by differential elutlon, and determina-
tion of the their radioactivity Indicates their relative affinities
for the ligand. In the placenta, where albumin and AFP are separated
by the placental membrane, these different affinities lead , in this
model t to the net diffusion of ligands toward the higher-affinity pro-
tein.

PROOF OF THE MODEL

We have used this assay to investigate the exchange of fatty
acids and diethylstilbestrol (DES)^ a widely used non-steroid estrogen
which has been shown to be a teratogen and transplacental carcinogen
(21^22), between bovine and human albumin and AFP. Our results^
reported elsewhere (23), confirm the preferential binding of PUFA by
AFP and support the AFP-mediated uptake of PUFA by the fetus. They
also indicate that DES» 1 ike PUFA, shows strongly preferential binding
to AFP. It appears from this that the albumln-AFP exchange system at
the placenta may extend to DBS and concentrate this drug in the fetus.

Another element oi ligand transport across the placenta Is

- 336 -

xndicaLed by our finding that PUFA compete strongly with DES for bind-
ing si tes on AFP. A specific implication of this is that PUFA levels
modulate DES binding to AFP and therefore DES uptake by the fetus.
More generally, since vacant binding sites on AFP are required for the
facilitated diffus ion mechanism to work, I Igand transport would seem
to be a function of the various competing ligands present in the fetal
circulat ion.

This finding of DES binding by mammalian AFP is slightly at
odds with the literature on the subject. Rat AFP has been shown to
bind DES (24) and DES transfer from maternal to fetal serum in the rat
has also been seen (25). Rodent and human AFPs, howver^ show dif-
ferent ligand-binding specificities (26), and DES binding to mammalian
AFP is less we 11 known. Sheehan and Voung (27) compared DES binding
in the plasmas of pregnant rats and humans (where AFP of fetal origin
is fotind In small amounts) with that of non-pregnant plasmas. Since
no elevat ion in DES binding was seen in pregnancy plasmas , they
inferred that AFP does not bind DES with high affinity. This
discrepantv with ptiblished data of DES binding to rat AFP and with our
results using human AFP is difficult to understand. It may reflect
competition from other ligands in the plasmas used, or differences
between AFP in the maternal and fetal circulation. At any event , DES
transfer into the fetal circulation has been seen in the rat (25) and,
tragically, in humans (28). In comparing our findings to those from

- 337 -

previ ous binding studies » it is inportant to note also that our assay
does not indicate what is conventionally thought of as high-affinity
binding, but rather relative binding affinity for the two proteins .

CONCLUSIONS AND IMPLICATIONS

AFP , 11 ke HbF, is most abundant during early gestation when
organogenesis occurs (albumin appears in the fetus later in gesta-
tion). This is consistent with a role of AFP in ensuring steady
nutrient extraction from the maternal bloodstream regardless of varia-
tions In the maternal nutritional state during the critical period of
development . This is also the period of greatest sensitivity to tera-
togens (29). This must be due in part to the rapid differentiation
and growth at this time, but is also consistent with AFP^mediated con-
centration in the fetus of teratogens with relatively high affinity
for AFP. Testing of more teratogens by our ConA-Sepharose assay will
cast more light on this hypothesis.

We see implications of these findings In two areas, namely the
elucidation of fetal nutrition and the screening of potential terato-
gens . In the former area, ligand (Including teratogen) uptake appears
to be modulated by the presence of competing 1 igands (e .g.PUFA) in the
fetal serum. The pharmacokinetics of this needs further Invest Iga-
tion. In the latter area, the assay quickly and quantitatively meas-
ures 1 igand binding to human albumin and AFP, and might allow

- 338 -

prediction of the extent of transport of a given llgand across the
plaiienta. As a step in teratogen screening, this would represenL a
great improvement on presently used in vivo tests, which ^ -f expen-
sive, time-consuming, and whose results, as shown by the case of
thalidomide ( 30) , not completely applicable to humans .

- 339 -

FIGURE 1

Model for cross-placental uptake of ligands by the fetus via exchange
between maternal and fetal proteins. It has been established that
fetal hemoglobin (HbF) has a higher affinity for oxygen than adult
hemoglobin (HbA), and that alpha-f etoprotein (AFP) has a higher affin-
ity for polyunsaturated fatty acids (PUFA) than human serum albumin
(HSA) . Facilitated diffusion of these ligands across the placenta
toward the higher affinity protein is generally accepted. Our results
show that AFP binds dlethystilbestrol (DES) more strongly than albu-
min, indicating that OES, a teratogen and cross-placental carcinogen,
is concentrated in the fetus by the same mechanism.

- 340 -

FIGURE 2

The ConA-Sepharose assay of the relative affinities of albumin and AFP
for a ligand (triangle) ALbumin, AFP, and a trace amount of radiola-
belled ligand are mixed and passed through a ConA-Sepharose affinity
column. An equilibrium is reached between free ligand (open trian-
gle), ligand specifically bound to albumin and AFP (closed triangles),
and ligand non-specif ically bound to the column matrix. This simu-
lates In vi^tro the equilibrium between maternal albumin, the placental
membranes, and fetal AFP 2:1} ^iyo. AFP is retained on the coumn due to
its sugar moiety (S) while albumin is eluted, separating the proteins.
(They are separated i^n vivo by the placental membranes.) Comparing the
radioactivity associated with each protein thus separated indicates
their relative affinities for the ligand.

- 341 -

FIGURE 3

Chromatogram of human albumin (peak b) and AFP (peak e) resolved from
a 10:1 albuminiAFP mixture by a ConA-Sepharoae affinity column. The
mixtures contained a trace amount of radlolabelled (A) docosahexae-
noate (C22:6) and (B) palmitate (C16:0). Most of the polyunsaturated
fatty acid (PUFA) C22:6 was associated with AFP, even in the presence
of a ten-fold excess of albumin, while most C16:0 was associated with
albumin. This confirms the preferential binding of PUFA by AFP.

FIGURE 4

Chromatogram similar to those in Fig. 3, but using radiolabelled
diethylstilbeatrol (DES) as ligsnd . DE5 , like C22 :6, shows strongly
preferential binding to AFP.

- 342 -

REFERENCES

1. Moore, K. L. The Developing Human - Clinically Oriented Embryology,
third edition, W, B. Saunders. Philadelphia, 1982

2. Hill, E.P. and Longo, L.D. Dynamics of maternal-fetal nutrient
transfer. Fed. Proc. 39 (1980) 239.

3. Huehns , E.R. Molecular changes in hemoglobins during development
and their functional significance, in Protides of the Biological
Fluids, H. Peeters, ed. Pergamon Press, New York.

4. Bauer. C.H.. Ludwig, I., Ludwig, H. Different effects of 2.3-
diphosphoglycerate and adenosine triphosphate on the oxygen affinity
of adult and fetal hemoglobin. Life Sci. 7 (1968) 1339

5 . Tyuma , I . and Shimizu, K. Different responses to organic phosphates
of human fetal and adult hemoglobins. Arch. Biochem. Biophys. 129
(1969) 404

6. Bellingham, A.J., Detter, J-C, and Lenf ant . C. Regulatory mechan-
isms of hemoglobin oxygen affinity in acidosis and alkalosis. J.

Clin. Invest. 50 (1971) 700

7. AsLrup, P., Rorth, M., and Thorshague, L. Dependency on acid-base
status of oxyhemoglobin dissociation and 2 , 3-diphosphoglycerate levels
in human erythrocytes. Scand. J. Clin. Lab. Invest. 26 (1970) 47

8. Wood, W.G. and Weatherall, D.J. Haemoglobin synthesis during human
foetal development. Nature 244 (1973) 162

9. Bard, H. Postnatal fetal and adult hemoglobins in early preterm
newborn infants. J. Clin. Invest. 52 (1975) 1789

10. Berde, C.B., Hudsone. B.S., Simoni, R.D., and Sklar, L.A. Human
serum albumin - spectroscopic studies of binding and proximity rela-
tionships for fatty acids and bilirubin. J. Biol. Chem. 254 (391)
1979

11. Muller, W.E. and U. Wollert. Human serum albumin as a "silent
receptor" for drugs and endogenous substances- Pharmacology 19 (1979)
59

12. Bassett, M., Defaye, G. , and E.M. Chambaz. Study of steroid-
protein interactions by electron spin resonance spectroscopy. Binding
of a SI) in- label led di hydro testosterone to bovine serum albumin.
Biochem. Biophys. Acta 491 (1977) 434

- 343 -

13 . Westphal , U. Steroid-Protein Interactions . Springer-Verlag, Ber-
lin, 1971. Chapters 6,7.

14. Sellers, K.M. and Koch-Weser , J. Clinical implications of drug-
albumin interaction, in Albumin Structure, Function, and Uses , V.M.
Rosenoer et al., eds. Pergamon Press, Oxford, 1977. p. 159

i5. Eiferman, A.E., Young, P.R., Scott, R.W., and Tilghman, S.M.
Intragenic amplification and divergence in the mouse alpha-f etoprotein
gene. Nature 294: 713

16. Kuoslahti, E. and Terry, W.D. Alpha-fetoprotein and serum albumin
show sequence homology. Nature 260 (1976) 604

17. Ruoslahti , E. and Engvall , E. Immunological cross reaction between
alpha-fetoprotein and albumin. Proc. Natl. Acad. Scl, USA 73 (1976)
4641

18. Hiral, H., Nishi, S., Watabe, H., and Tsukada, Y. Some chemical,
experimental , and clinical investigations of alpha-fetoprotein. Gann.
Monogr. Cancer Res. 1^4 (1973) 19

19. Parmelee, D.C., Evenson, M.A., and Deutsch, H.F. The presence of
fatty acids In human alpha-fetoprotein. J. Biol. Chem. 253 (1978)
2114

20. Ruoslahti, E. , Estes, T. , and Seppala, H. Binding of bilirubin by
bovine and human alpha-fetoprotein. Biochem. Blophys. Acta 578 (1979)
511

21. Stillman, R.J. Am. J. Obstet. Gynecol. U2 (1982) 905

22. Herbst, A.L. Obstet. Gynecol. Annu . 1^0 (1981) 167

Hsia, J. C, Wong, L. T. , and Deutsch, H. F. manuscript submitted for
publ icat ion.

24. Savu, I.. , Renassayag, C. , Vallette, G, and Nunez, E.A. Ligand
Properties of diethylstilbestrol : studies with purified native and
fatty acid-free rat a lpha-1 -fetoprotein and albumin. Steroids 34
(1979) 737

25. LeGuern, A., Benassayag, C, and Nunez, E.A. Role of alpha-1-
fetoprotein in the transplacental transfer of natural and synthetic
estrogens in the rat. Dev. Pharmacol. Ther. 4:8uppl. I (1982) 79

26. Aussel, C. and Masseyeff, R. Comparative binding properties of rat

- 344 -

and human alpha-f etoproteins for arachidonic acid and estradiol. R'ls.
Comm. Chem. Pathol. Pharmacol. 42 (1983) 261

27. Sheehan, D.M. and Young, M. Diethyls tilbestrol and estradiol bind-
ing to rat serum albumin and pregnancy plasma of rat and human. Endo-
crinology 1^04 (1979) 1442

28. Herbst, A.L., Ulfelder, H., and Poskanzer, D.C. New Engl. J. Med.
284 (1971) 878

29. Wilson, .I.G. Critique of current methods for teratogenicity test-
ing in animals and suggestions for their improvement , in Methods for
detection of environmental agents that produce congenital defects ,
T.H. Shepard at al., eds . North-Holland Publishing Co., Amsterdam,

1975.

30. McBride, W.G. Lancet 2 (1961) 1358

CROSS-PLACENTAL UPTAKE

MATERNAL
CIRCULATION

PLACENTA

FETAL
CIRCULATION

Hb-A —

- o, ^

^ Hb-F

HSA —

^ PUFA _
" (DES) *

T AFP

FiauR£ I

PUFA
(DES)

in vitro MODEL OF C ROSS-PL ACENTAL UPTAKE

F } c u RB Z

\ >

f\i

Z
CT

m
o

CD
(E

(A) C22:6

7 -

6 -

Z. 5 -

4 -

3 -

a r

I -a ; —

3.2

i?.4

- 1 .6

a
o

- .e

^ 1 C.ii

E
C

K
O

in
m
d

- 3

^ 1

73
H

a
i-i
o

-sJ

VOLUME CML)

in vitro MODEL OF C ROSS-PL ACENTAL UPTAKE

F I a u f^c 3

- 348 -

RRDIOflCTIVITY x 10"^ (

-3 r__.

)

■■ m

U
71

-J
O
>

u) m

) 01 X uju 0B2 lb 3DNda^0Sati

RRDIORCTIVITY x 10"^ ( "- ")

- tn

_j
o
>

u) m

n w

( ) 01 X UJU 092 j.y 33NdHyOSHH

- 349 -

EFFECT ON THE TISSUE OF YOUNG FISH AND RATS
OF EXPOSURE TO
LEAD, CADMIUM AND MERCURY

I- M. NichoUs, K. Teichert-Kuliszewska and M.J. Kullszewski. Department
ni lUo'iogy, York University, Downsview. Ontario. M3J 1P3

and
u.R. Girsis. Department of Biological Science and Applied Chemistry. Seneca
College, WilJowdale, Ontario. M2J 2X5

- 350 -

TABLE OF CONTENTS

Page

Introduction 1

Background 1

i. Rat Kidney Enzymes 3

Animals 3

Lead Analysis 5

Results 6

Discussion 8
Tables 1, 2, 3 and Figure 1

II. Rat Kidney mRNA 10
Animals 10
Results 12
Discussion 14
Figures 1, 2. 3

III. Muscle 16
Figures 1, 2, 3

IV. Fish Enzymes 17
Animals 17
Results 18
Tables 1,2,3 ,^*-

V. Summary 18

References ^9

Acknowledgement 21

- 351 -

ABSTRACT

EFFECTS OF EXPOSURE TO LEAD. CADMIUM AND MERCURY ON THE TISSUE OF YOUNG
FISH AND RATS

P.M. Nlcholls , K. Teichert-Kullszewska and M.J, Kullszewskl. Department
of Biology, York University, Downsview, Ontario, M3J IP3

and
G.R, Girgis, Department of Biological Science and Applied Chemistry, Seneca
College, Willowdale, Ontario, M2J 2X5

The response of the kidney and liver, both in mammals and fish exposed
to cadmium and mercury involves the stimulation of transcription of the
genes to provide messenger RNA (mRNA) for metallothioneln species. However
recent work shows that neither tissue metal levels nor tissue metallothioneln
levels Indicate functional Impairment or functional reserves In the tissue.
The sltution Is even more complicated with Pb^"*" exposure since there Is no
triggering of metallothioneln gene transcription. However we have found
that (1) the livers of rats do have Increased messenger RNA for certain
proteins which are released into the blood and (2) the kidneys of rats do
have increased mRNA for a prominent protein-splitting enzyme, urokinase.

In addition to these different genetic responses to Cd^"*", Hg^"*" and Pb^"*",
the enzyme function at the kidney membrane/urine interface (I.e. the brush
border membrane) has been studied in rats. Either an acute dose of Pb2+
(10 mg Pb2+/kg body weight) or chronic administration of Pb2+ In acidified
drinking water (500 mg/1) for 3 or 6 weeks exert effects on the brush border
membranes which are not correlated well with the Pb2+ level In the fraction
and which are not accompanied by detectable changes in the membrane or in
the urinary function. The same was true after an acute dose of Cd2+. The
effects on enzyme activity result from some mechanism not related to a
direct effect of the heavy metal and the effects would be undetected if the
animals were not sacrificed for the analysis.

Thus we have analyzed the muscle of Pb2+ exposed animals to see if a
more accessible tissue could provide evidence for heavy metal exposure.
The mRNA obtained from the muscles of rats given Pb^"*" alone, Cd2+ alone, or
Pb^"*" plus Cd2+, all showed increased activity, greatest in the Pb2+ + Cd^^
experiment, especially for a protein believed to be glyceraldehyde-3-phosphate
dehydrogenase. This is an enzyme affected by heavy metals in vitr o , and thus
the gene transcription product, mRNA, may be produced in Increased amounts as
a compensatory response.

In studying the muscle of a different animal, namely fish, exposed to
acidified water containing 100 yg Pb2+/1 or 500 pg Pb2+/1 this enzyme
exhibited decreased activity. The liver and gonads of the fish also
exhibited changed activities for certain enzymes. A much lower Pb^"*" level
in the water (5 ug Pb^"*"/!) has been used for much longer periods (5 months
rather than 3-6 weeks) in order to see what minimal levels of Pb^"*" could
be tolerated with no development of changes In these very sensitive para-
meters of tissue function.

- 352 -

INTRODUCTION
The goal of the research is to study the effects of various
levels of lead, cadmium and mercury presented to mammals and fish
in acidified water. The enzymes and the messenger RNA of muscle,
liver, kidney and gonads are being studied with a view to providing
a broad data-base for dose-response and to developing a biomonitor
test. A number of recent studies have established that blood and
tissue metal levels are poorly correlated with functional Impair-
ment. The approach we have taken has been:

(1) either to administer one to three doses of a low level of
these compounds in order to study acute effects, or

(2) to present the heavy metal In the food or in the water at
various concentrations for a number of weeks to study
chronic ef f ects .

We have compared protein synthesis and a number of enzymes in detail
in the liver, kidney, muscle and gonads.

BACKGROUND

In a recent study (l) of wild and captive birds and laboratory
mice, it was found that serious kidney pathology occurred at Cd
concentrations of the same order as found in western adult human
populations (^200 mg/kg wet weight cortex), a level considered by
some authorities to have no serious effects on kidney function
which is a major target of Cd toxicity (1, 2).

The effect of Hg alone was not so clear in this study of wild

- 355 -

anJ captive species because in the environment there is biotrans-
formation between inorganic and organic forms of Hg, especially
methylmercury. Recently, It was reported that the proportion of
methylmercury to total mercury in river waters in Canada was 30% (3) .
Both forms of Hg are known to be extremely toxic In mammals and
fish, with the major targets being nervous tissue and liver for
methylmercury and kidney and liver for inorganic Hg.

We and others have obtained some preliminary evidence that
cadmium, lead, and also methylmercury induce synthesis by the
liver of specific serum proteins known as "acute phase reactants" (A-7) .
In rats, the most abundant and readily detected of these is a , -
acid glycoprotein , Mammalian liver responds to various injuries
such as acute inflammation, infection, etc., by profound changes
in the gene activity providing messenger RNA (mRNA) for these
"acute phase reactants".

While current research is directed towards uncovering knowledge
of their gene control, less is known about the function these

serve .

A second important group of genes respond to Hg and Cd , but
apparently not to lead, are those coding for the different isomolecular
forms of metallothionein(8) . We have recently compared expression of
these genes in liver and kidney following a single small dose of
Hg^"^ (0.5 mg/kg or 1.0 mg/kg) and find a striking difference between
the tissues. Kidney exhibited a large and statistically significant
response only at the lower dose and liver exhibited a statistically

- 354 -

significant response only at the higher dose. Such remarkable
sensitivities in gene expression to a single heavy metal do
underline the need for a wider number of carefully prepared
laboratory test situations.

Other criticisms of the sole use of metallothionein-bound
heavy metals as an indicator of toxicity are found in recent
reports that metallothionein I, present in various proportions of
the total, depending on age and species, is very labile during
handling and preparation from tissues, relative to the more adult-
prevalent form, metallothionein 11(9). Other difficulties in inter-
pretation result from a recent report showing that heavy metal can
bind to low molecular weight proteins isolated in the metallothionein
fraction but quite distinct from metallothionein ( 10) .

Thus, neither tissue metal levels nor levels of metallothionein
are indicative of functional impairment or functional reserves.

I . Rat Kidney Enzymes

A nima ls

Male rats (125-150 g) of the Wistar strain (Woodlyn Laboratory,
Ltd., Guelph, Ontario, Canada) were fed on Purina 5001 Rodent
Laboratory Chow and water ad libitum . 2.5 g of lead acetate
(Pb(CH COO) .3 HO), were disolved in 2.5 litres distilled water

J ^ ^

(i.e. 0.1% lead acetate) to which 1.0 ml of concentrated HCi. was
added to preclude the precipitation of insoluble lead salts as
described previously (Aungst et al. , 11 ; Michaelson and Bradbury,

- 355 -

12 ). Control animals received Cap water containing a similar
amount of acid. This regime was continued for three weeks or
for six weeks, at which time the animals were removed to metabolic
cages where they were kept for 2^ hours. The urine was collected
in iced containers and was analysed directly or kept frozen. The
animals were killed and the kidneys removed and placed in ice-cold

containers. Other rats received two consecutive daily ip

2+
injections, each of 0.546 mg Pb /lOO g body weight, as described

previously and were killed 24 hours later (Nicholls et al . , 13 ) .

The subsequent steps for preparation of the brush border were

carried out in an ice-cold container.

The enzyme activities of the brush border were expressed as
umoles (product formed) /mg proteln/hr and of the urine, umoles
(product formed) /hr/mg creatinine excreted as suggested by Bonner
e t a 1 ■ , (14 ). For the urine, the number of animals studied at
3 weeks was 19 control and 23 treated while the number studied at
5 weeks was 18 control and 24 treated. In each case 3 separate
experiments were carried out. For the brush border three or four
separate preparations were studied; each one of these preparations
consisted of 6-8 control rats and 6-8 treated rats .

The preparation was based on a modification of the method of
Liang and Sacktor (15 ) and had negligible mitochondrial contamination
The kidney homogenate was prepared in 3 vol of 0.25 M sucrose made
up in 10 mM Tris. HCl (pH 7.5 at 20°C) . After repeated centri-
fugation steps the preparation was placed over a continuous sucrose

- 356 -

j'ladiuiii rrom 32.5'/', (w/w) to 41.5% (w/w) made up in the Tris buffer.
These tubes were centrifuged at 100,000 g for 1 hr at 4 C. Following
puncture of the tube, twenty fractions of 0.5 ml were collected,
starting from the bottom of the centrifuge tube.

The enzymes were measured as described previously (Nicholls
et al., 16 ) and were y-glutaniyl transpeptidase, alanine amino-
peptidase, alkaline phosphatase, and (Na + K ) adenosine triphos-
phatase and urokinase. The latter two enzymes were located in the
basal-lateral membrane fractions 10-20 that were also collected.
The urokinase assay was a modification of the method of Highsmith

( 17 )■

Lead Ana lysis

The concentrations of lead in the brush border fractions were
determined in acid-washed glassware by the digestion of 1 ml tissue
preparation with 0.5 ml HNO and 0.3 ml H2SO, . The resultant
clear solution was diluted with double distilled water and 0.010 ml
was injected into the Scintrex AAZ-Z Zeeman modulated atomic
absorption spectrophotometer. Lead standards were prepared in the
appropriate matrix. The detection limit was less than 2 ng/ml ,
the rcciwery of Pb^ was better than 95% and the standard error of

inuitiplu determinations was + 8.5%. Each preparation measured was

2+ 2+

derived from 3 control rats and 3 Pb exposed rats. No Pb

2+
was detected in control homogenates ; Pb measured in control

mumbrane fractions, owing to difficulty in acid-washing the ultra-

centrifuge tubes, were subtracted from Pb values for dosed animals

- 357 -

700-

350 -

FIGUTtE LEGEND
Fig. 1. The activity of enzymea in the brush border fractions collected from
the sucr,i3e gradient centrlfugatlon step. Panels A, B and C were from controls
and Jr-im animals exposed to 0.1% lead aceUte In the drinking water for 3 weeks,
Panels [', E, and F were frnm controls and from animals exposed to lead for 6
weeks. A and n, >-glutainyl transpeptidase; B and E, alanine aminopeptldase; C
«n.j F. alkaline phosphatase. Enzymes activities are expressed as umoles/mg
pr:iteln/hr),

o, control

•. lead treated

- 358 -

RESULTS
For the rats that were exposed to 0.1% lead acetate in their
drinking water for three weeks changes in the specific activities
of the three brush border enzymes were measured (Fig. iA,B,C).
The three peak fractions for each enzyme were combined and the
specific activities compared for control and treated animals. In

the case of y-glutamyl transpeptidase and alanine aminopeptidase

2+
the activities of Pb -treated rats exhibited decreases that were

statistically significant (P<0.05). The increase in activity of

alkaline phosphatase was statistically significant (P<0.05)

although the magnitude was less than that of the other two enzymes.

2+
When rats were exposed to Pb for six weeks and the brush

border fractions analyzed (Fig. ID,E,F) there were similar changes

to those seen at three weeks. The activities of the peak fractions

2+
also resembled those after three weeks of Pb exposure in that

, -glutamyl transpeptidase and alanine aminopeptidase were significantly

reduced while alkaline phosphatase activity was increased. The

latter change was not statistically significant, however, after

six weeks exposure.

The activity of urokinase was greatest in fractions 10-12

and the activity of (Na + K ) -ATPase was greatest in fractions

1 2- 1 A . Fractions 10-20 are believed to be the basal-lateral

membrane fractions of the tubular cells. There was no change in

l\\e activity of urokinase in these fractions derived from the rats

2+
exposed to Pb" for three or six weeks (6. 79 ± 0.96 umoles/mg

- 359 -

proteln/hr for controls compared to 5.A9 + 0.75 umoles/mg protein/

9+ + +

hr for Pb exposed rats). Similarly, (Na + K ) -ATPase did not

2+
exhibit altered activity in these fractions after chronic Pb

t'xposure .

2+
The effect of Pb in the drinking water in these subacute

2+
oral administration experiments was similar to the effect of Pb

In the previous acute injection experiments, except for reduced

effects in the case of alkaline phosphatase actlvlty05)To see

2+
whether other parameters were also similar, Pb was administered

by acute injection or by the subacute oral route for 3 or 6 weeks.

The brush border fractions were examined for protein, phospholipid

2+
and Pb content (Table 1) as well as for microscopic changes.

The protein contents and phospholipid contents were unchanged

2+ 2+

either by acute injected Fb or by subacute Pb treatment. The

proportion of phospholipid (26%) is similar to that reported for

rabbit kidney brush border. No detectable change following

2+
Pb administration was seen in an electron microscopic examination

of the brush border membrane.

2+
The homogenate Pb levels at 3 weeks were similar to those

reported at 2 weeks in similar experiments of Aungst et _al . , ( 11 )

with a blood level of 20 ug/dl. In treated rats the brush border

2+
Pb level was greatest in the acute injection experiment, 21 ng/g

2+
kidney, which was 0.4% of the Pb in the total homogenate. In the

subacute oral experiments the level was twice as high (8.1 ng/g

kidney) after 6 weeks' exposure as after 3 weeks' exposure

- 360 -

2+
(3.1 ng/g kidney), values which were 0.2% and 0.1% of Pb levels

measured in the total homogenates , respectively . It was thus

2+
possible to calculate the level of Pb contributed to each of the

enzyme assays by the addition of the brush border membrane p''t"eln.

Vnr uxampie, the level for the 6 week experiment was approximately

-9 2+
3.3X10 MPb for the y~ glutamyl transpeptidase assay , and

— 8 2+
approximately 1.5X10 M Pb for the alanine aminopeptldase and

2+
alkaline phosphatase assays. It is unlikely that these Pb levels

2+
affected Che enzyme activities, however, since adding Pb directly

to assay tubes containing brush border from control kidneys (which

contributed endogenous Pb of 8 X 10 M) had no effect (Table 2).

Tab] e 3 shows the activity of the enzymes found in the urine

2+
of animals exposed to Pb for six weeks. There was no significant

2+
effect of the Pb exposure detectable in urine volume, protein,

creatinine excretion or enzyme excretion. After the shorter

exposure the results were similar except that there was a small but

statistically significant decrease in urine volume. It is thought

Chat the decrease in urine volume reflects a decreased fluid intake

2+
based on aversion to Pb described by others where higher doses

were administered (Mlchaelson, 18 ).

DISCUSSION
The present observations of the altered activiciy of 3 enzymes

located in the brush border of the kidney of rats following subacute

2+
exposure to orally administered Pb are similar to Chose following

- 361 -

Table 1

2+
Effect of acute or subacute Pb exposure on

2+ a
kidney brush border protein, phospholipid and Pb content

BRUSH BORDER FRACTIONS

j_^ Protein Content

Time of Pb (mg/g Kidney) Lead Content

Kxposure Control L ead ( ng/g Kidney )

2 days 0.157 ± 0.03 0.160 ± 0.02 21.1 ± 2.92

(A) (4) (3)

i weeks 0.110 + 0.01 0.135 ± 0.02 3.11 + 0.43

(4) (4) (2)

6 weeks 0.155 ± 0.01 0.125 ± 0.01 8.08 ± 0.99

(4) (4) (2)

Phospholipld/Protein (mg/mg)
2 days 0.26 0.24

6 weeks 0.28 0.26

The analyses were carried out on the brush border fractions containing
the peak enzyme activities, x * SE. Number of preparations shown in
parentheses, each of which was derived from 3 to 6 rats.

- 362 -

Table 2
Absence of effect of Fb added to kidney brush border fractions in vitro

(

Enzyme Activity
^umoles/mg protein/hr)

rnnc. I
(M)

Added Y-Glutamyl
transpeptidase

Alanine
amino peptidase

Alltaline
phosphatase

None
(endogi
( 8 X

2nous Pb )
10-10 ^

+
(8)

37.1

93.6

+ 6.96
(10)

108 + 8.80
(10)

10-^

i+71

(6)

hl,k

87.8

+ 8.56

(8)

97.6 + 7.73

(8)

10-^

46i

+
(6)

25.2

89. i^

+ 8.30
(8)

97.8 + 7.^0
(8)

IT'^

U80

+
(2)

20.0

91.2

+ 2.80
(2)

133 + 2 . 50
(2)

rr^^

+
(8)

23.0

91.^

+ 6.80

(10)

106 + 8.63
(10)

!0 ■

50U

+
(2)

8.5

91.1

+ 9.00

(2)

lUi+ + k. 00
(2)

lo-'^"

526

+
(8)

35.1

89.3

+ 6.60
(10)

113 + 7.17
(10)

'^ Each brx-ish border preparation was from 6-8 rats, x + SE for number of
r.ramples tested ,

- 363 -

Table 3
Urinary protein, volume, and creatinine and the
activity of enzymes from rats receiving
lead acetate in the drinking water for 6 weeks

Control Lead

Protein (nig/2l+ hr) 35,2 ± 1|.33 Uo.2 ± 3.06

Urine vol (ml/2i+ hr) 7,54 ± O.96 9.26 ± 0.88

Creatinine (mg/lOO g/2i+ hr) 1.68 ± 0.1? 2.12 ± 0, 18

Enzymes

(uuioles/hr/mg creatinine)

Y-Glutamyl transpeptidase 3^.0 ± U.98 3^.7 ± h.OS

Alanine aminopeptidase 12,5 ± 1. Bit 9.51 ± 0.99

Alkaline phosphatase 315 ±30.5 307 ±23.2

Urokinase ikj ± 17,7 ^ 13^ ± 16.8

The urine collection was in ice-cold flasks for 2k hr. T^e protein content

Hnd enzyme activity were assayed as for homogenate fractions, x ± SE for 18

contr-1 rats and 2k treated rats studied in 3 separate experiments. Student's t

test showed no statistically significant differences in means for lead-treated

rjits roitipared to control rats (i.e. all the P values were greater than 0.05).

- 364 -

2+
ihe Hcute i.p. administration of 1.0 mg Pb / 100 g body weight

(Niitiolls L-t ii\ . , 13 ). In die acute experiments, however, the

clianges were more marked in the case of alkaline phosphatase ,

wliich is the only enzyme to exhibit increased rather than deceased

2+ 2+

activities after Pb . Since the Pb content of the brush border

membrane fraction is highest in the injection experiments and is

higher after 6 weeks than after 3 weeks in the oral experiments,

there is thus no quantitative correlation of lead levels with

2+
changed enzymatic activities. Moreover, a direct effect of Pb

on the enzymes is unlikely in view of the absence of any effect of

2+
Pb added directly to membrane preparations at levels estimated

2+
to be present. Concentrations of Pb high enough to affect these

-3
enzymatic activities, reportedly 10 M (Vallee and Ulmer, 19 ),

would likely only be present in the nucleus and cytoplasm where

Pb .is localized (Barltrop e t al . , 20 ; Mistry et_al . , 21 )•

2+
In previous experiments with Cd in vitr o only aminopeptidase was

decreased at 10 M Cd while with Pb , none of the enzymes was

-A
changed at 10 M (Nicholls e^_al . . 13 ; 16 ).

Tlitf appearance of enzymuria in rats following the effects of

2+
lip on the brush border membrane (Kempson et^_al . , 22 ; Nicholls

('[ _.'il . , 1 (S ) as well as following other nephrotoxic substances,

can provide a non-invasive test of renal integrity (Price, 23 ) •

2+
However , in the case of Pb exposure no increased enzymuria or

damaged brush border can be detected in the acute or subacute

2+
stages of Pb intoxication at least with relatively modest amounts

- 365 -

2+
of Pb . No changes in protein excretion or other urine functions

2+
could be detected after the Introduction of Pb (0.05%) In the

drinking water for 3 or 6 weeks. In spite of these negative

findings, there were readily measured changes In the activity of

three important brush border enzymes. No such change was found

In the activity of another enzyme that was measured (the Na + K

ATPase) , and this is believed to be In the basolateral membranes

(Kyte, 24 ). The location and function of these enzymes have

been discussed previously (Nlcholls etal, , 13) .

The present experiments, directed towards the basement membrane

preparations demonstrate that statistically significant changes in

the activity of the membrane enzymes occur after acute Ip Pb or

2+
subacute oral Pb exposure. These changes, which are subtle and

cannot be readily detected by microscopic or conventional urinary

enzyme measurements, nonetheless must be considered in attempts to

2+
understand the effect of Pb on kidney.

II. Rat Kidney mRNA
Animals

Lead acetate was administered in one intraperitoneal injection
of 1 mg/100 g body weight In 0.2 ml (i.e. 0.546 mg Pb /lOO g body
weight). Control rats received the same volume of water (0.2
ml/100 g body weight). The kidneys were obtained 48 h later and

- 366 -

were pooled either from 6 control and 6 treated rats (mRNA Isolation)
or from 2 control and 2 treated rats (homogenate studies) . The
kidneys were rapidly removed and chilled in an iced container and
all subsequent steps were carried out in the cold. The tls&uc: was
homogenized in 3 volumes of buffer 1 containing 0.05M Tris/HCl
(pH 7.5) and O.IM NaCl together with 0.25M sucrose. Messenger RNA
was obtained by deproteinization of the rlbosomal fraction and
purified by oligo (dT) cellulose chromatography. This mRNA was
measured for translational activity in a rabbit reticulocyte
protein synthesizing system.

The products of translation contained in 25 ul of the reaction
mixture were analyzed by electrophoresis in cylindrical polyacry-
lamlde gels (5 mm diam x 100 mm long) containing SDS as described
before (25) , The discontinuous system consisted of a 10% (w/v)
acrylamide gel, pH 8.8 on which a 3% (w/v) polyacrylamide spacer
gel, pH 6.8, was layered. Samples and standards were electrophoresed
for h hr at 22 C at a constant current of 2.5 mA/gel, Following
fixation in 50% trichloroacetic acids, the gels were sliced (2 mm),
solubilized and counted as described above. In other experiments
isoelectric focussing was carried out similarly in cylindrical
gels, after which the gels were placed on polyacrylamide slab gels
and subjected to electrophoresis as described by O'Farrell (26).
The gels were stained with Coomassie brilliant blue, destained by
diffusion in 7.5% (v/v) acetic acid, dried, and subject to radio-
autography using Kodak no screen X-ray film.

- 367 -

The urokinase assay was a modification of the method of
Highsmlth (17) and the fibrinolytic assay was carried out as
described by Johnson et al . , ( 27 ) «

RESULTS
Fig. 1 shows the release of tyrosine from casein by digestion
with plasminogen that had been activated by kidney homogenate (i.e.
urokinase activity) . There was a linear increase with incubation

time from 10 to 60 mln using control kidney. With kidney homo-

2+
genate from the Pb -treated rats, the urokinase activity was

significantly higher and was linear from 20 to 60 min. The

addition of increased amounts of homogenate protein up to about

3 mg from control rats increased ot-casein digestion but larger

amounts had little further effect. When homogenate protein from

2+
Pb -treated rats was used and compared to that from controls, a

significant increase in a-casein digestion was detected at each
point up to 5 mg homogenate protein. When measured in the fibrino-
lytic assay, the plasminogen activator activity of the homogenates

2+
from the Pb - treated rats was twice that of the control rats,

thus confirming the results of the a-caseln digestion assay.

Since the urokinase activity of the mitochondrial- lysosomal

fraction (P15) was higher in kidney preparations from Pb -treated

rats than from control rats just as In kidney preparations treated

with a detergent, it seemed possible that the presence of Pb in

the kidney might be acting directly on membranee to release

- 368 -

2 0V

% ■/

! /

20 40 60
TIME tmin)

12 3 4 5
PROTEIN Imq)

Fig 1. Urokinase ;ictivity of kidney homogenates from
control rais and from rats followmg the injection of lead
acetate. (A) Incubation of 2 mg protein with plasminogen
for various times prior to the addition of tt-casein. as
described under Materials and Methods. (B) Incubation
tor 30 mm of various amounts of protein as described under
Matcriiils and Methods. Points plotted arc mean ± S.E.M,
for four preparations Key: (Ol control; and (•) lead.

^ 0.5

o fo-s la^ 10"*

LEAD ACETATE (M)

Fig. 2. Urokinase activity of kidney homogenates from
control rats following the addition of lead acetate m ciiro.
Kidney homogenate (2 mg) was incubated for 30 min as
described under Materials and Methods. Lead acetate was
added to the homogenate in the molar concentrations
indicated. Bars arc means = S.E.M. for four preparations

- 369 -

urokinase activity. With this in mind, Pb was added directly to

kidney homogenates obtained from control rats (Fig. 2). The addition

— 8 2+
of 10 M Pb did not affect urokinase activity. However higher

concentrations (10~ M Pb and 10 M Pb ) caused statistically

significant (P<0.05) decreases in urokinase activity. Thus the

increased activity seen in homogenates from animals treated with Pb

2+
In vivo are unlikely to arise from Pb ions acting directly on the

enzyme, in view of the opposite direction of the effect when Pb

was added in vitro .

Poly (A) mRNA was obtained from the kidneys of control rats

and of rats injected with lead and translated in a rabbit reticulocyte

-14 -
lysate using L Cj leucine as a precursor of proteins. The newly

labelled peptide products of the translation assay were separated

first by isoelectric focussing gels (Fig. 3A) and then by SDS

polyacrylamlde slab gels (Fig. 3B) . In the former, It can be seen

that there are 8 major groups of labelled proteins by radioautography ,

the most basic proteins being of isoelectric points of approximately

pH 9.0, 8,6, and 8,0.

When the Isoelectric focussing gels were sliced and counted
the preparations from the lead-treated rats were 80% more radio-
active for protein of pi 8.6 (i.e. 1050 dpm for control and 1810
dpm for treated) and 60% more radioactive for protein of pi 8.0.
The other areas were less than 40% more radioactive on preparations
following lead treatment.

Isoelectric focussing gels similar to these were subjected to

- 370 -

8.6

I Ml

3.5

I \ I

•

'..^U?W!lWt-w.

I rti

-- #

_ 68

'_47
~ 43

_ 23

_ 14

. #

_ 47
- 43

Fill * NoclcLlnc fouuMiig and SDS jiel L-k-ciri)phorcMs followed by radio;uiloariiph\ ol princins
s\mhcM/fd from ["C'llcucinc in mRNA-dcpcndoni rcticulocyic lys:iics m response lo kidncv polv(A)'
RNA Pol>[ Al' RNA i^ ug) Irom conlrol or irciiied animals was inciibiitod in 25 »! lysiUc L-oniiiimnu
I 'C|leuL-inc js dfsLTEhcd in Maicnals and Methods and ihe lysiite sub|ceied to (A) isi.eiceine foeusini;
or (H) isoelecliiL toeusmi; followed b\ SOS iiel electrophoresis, Conlrol lysatc (upper tiel A .ind upper
eel lileonl.iined ll).IHM)dpm Lvsiile Irom lead-Ireaied animals (hmer pel A ,md lower tci BleonLiined
ISJKHldpin Ihe p)f anti radioactiviiv of ihe iiel slices following isoeleclric focusing was deiernniied
on similar >icK .-Her soakinj: in distilled water as deseribed previously [S|. The molecular weijihls (A/,)
tollowin^; SnS i!e! cleelophoresis in ihe second dimension were deiermined iisnii; marker proteins,
hoMiie scrum .ilbumin ((iK.(HKI|. human urokinase (47.IHHIJ. ovalbumin |43.S(K)|, lr\psin (Z.v.MHl) and
hso/vme ( N.MHI) Fvposure Imie al -70" was fi hr (A) and I4hr(13|

- 371 -

SDS gel electrophoresis in the second dimension and radloautographed.
Many labelled peptides from the isoelectric focussing gels were
removed during SDS gel electrophoresis because of small size. Since
fluorography and long exposure times were not used, the labelled
products (approx. 50) of only the most abundant poly (A) mRNA
species were detected. The poly (A) mRNA directed relatively major
amounts of incorporation of i Cj leucine into 2 proteins of iso-
electric point and mol. wt of 8.6 and 45,000 respectively, and of

8.0 and 32,000 respectively. These 2 proteins showed increased

2+
density by radioautography in samples derived from Pb -treated

rats and possibly are the major 2 forms of urokinase such as

described in other animal species (28 ) • Because the amount of

protein produced in these experiments is extremely low, it is not

possible to perform a direct urokinase assay on the gels.

When marker proteins were subjected to the same electrophoretic

procedures, human urokinase had a pi and mol. wt of 8.6 and

approximately 47,000 and a second band exhibiting a slightly lower

mol. wt. When the SDS gel electrophoresis was carried out in tube

gels and the 45,000 dalton band was eluted and used as a source of

protein for isoelectric focussing gels, the dpm per slice for

2+
proteins of pi 8.6-9.0 was 1000 dpm using preparations from Pb -

treated rats compared to 500 dpm using preparations from control rats

DISCUSSION
The activity of the plasminogen activator in kidney homogenates

was measured by a-casein digestion and by fibrinolysis. Preparations

- 372 -

from control rats and rats that had received on ip Injection of
Pb^"*" (0.5 mg Pb^"'"/100 g body weight) 48 h prior to the experiment
were used and there was a marked increase in the activity in the
Pb ^"''-treated animals. Several observations suggest that this
proteolytic activity was not due to an increase in non-specific
proteolytic enzymes but rather was due specifically to an increase
in the plasminogen activator. For example, the assays used were

dependent upon the presence of plasminogen. Moreover, the poly (A)

2+
mRNA that was obtained from the Pb -treated animals was more

active in overall translation and specifically, in directing

the incorporation of [^^c] leucine into a protein of an approximate

molecular weight of 50,000 and pi 8.6. Increased incorporation was

also found with crude kidney homogenate preparations and this

would not be the case were increased amounts of various proteolytic

enzymes present in these preparations, since newly synthesized

peptides thus would be hydrolyzed rapidly. The location of the

greatest plasminogen activator activity in the lysosomal-mitochondrial

fraction resembled that reported by Maclag et al . , (28) for rabbit

kidney and cultured pig kidney.

2+
The presence of a cytosolic Pb -binding protein of mol. wt

63,000 (21) and the presence of lead-induced intranuclear inclusion

proteins (mol. wt 32,000 and pi 6.3) (29) appear unrelated to the

protein described here of mol. wt 45,000 and pi 8.6. The events

2+
resulting from a single exposure to a low dose of Pb could be

triggered e.g. by Pb directly activating a serine protease such

- 373 -

as plasminogen activator. This result seems not to occur since

Pb^^ added directly to tubes decreased plasminogen activity (Fig. 2).

Similarly, Pb^"^ might directly activate mRNA, but if so. It would

be expected that the mRNA would yield a number of products rather

2+
than one major protein. If, e.g. Pb stimulated the release of

2+
calcitonin through Its interaction with Ca and phosphate levels,

then this might possibly explain the Increased plasminogen activator

activity (30). In any case, not only are some brush border enzymes

reduced In their activity but some enzymes, such as alkaline

phosphatase of brush border and plasminogen activator of

membrane fractions are Increased in activity in the kidney of rats

2+
that have received one injection of Pb (0.5 mg/100 g body weight).

III. Muscle
These experiments were carried out on rats that received single

94- 2+

injections of either 0.1 mg Cd /lOO g body weight or 0.5 mg Pb /
100 g body weight. Other rats received a combination of Pb and Cd
at the same dose level. The rats were killed 48 hr later and thigh
muscle obtained for preparation of mRNA as described previously.

2+
This dose of Cd had previously been shown to stimulate the

cranslational activity of kidney metallothionein mRNA while this

dose of Pb , as seen in the previous section, stimulated kidney

urokinase mRNA. As can be seen In Fig. 1, 2, and 3, the translational

activity of the mRNA obtained from muscle was also stimulated. The

protein products that showed increased labelling were of Mr similar

- 374 -

Fir- 1.2,3- 3DS pel electrophoresis and isoelectric focusing
■>': proUMnr, s.ynthpsized f rom \"g\ leucine in a mRNA -dependent
:..l.i(:ulocyt<> lysato. (a) [Vluscle poly (Af RNA was added to the
iyr-^i.t.r contained in 50 ul reaction mixturp. After incubation
ior 1 hr at 26 C, 25 ul of the mixture war. subjected to oDS
polyacrylamide gel electrophoresis, sliced and the labelled
Drotoins solubilized and counted. Marker proteins were actin (A),
tropomyosin (T), glyceraldehyde-3-phospH<iedehydrogenase (G).
and lyc^ozyme (L). (b) Isoelectric focusinfr of protein obtained

from ?d.icos 19-2?. obtained from gels prepared as described
undfr (a), .^he pH and incorporation of V'^'cUeucine into protein
wpre measurod on gel slices. Control (o) ; i'.xposed (•)
y^ir.l Cd ; Fip. 2.Pb ; Fig. 3 Cd+Pb.

Fig. 1 Control Oi Cadmium • - 375 -

Fig, 2 Control 0; Lead •

_l
(/)

E
a.

Fig. 3 Control 0| Cadmium+Lead •

- 376 -

to actin (Mr 45 ,000) with two discrete protein peaks. The slices
19-27 from the SDS polyacrylamide gel electrophoresis fractionation
step were eluted and subjected to isoelectric focussing in gels.
The protein from the i peaks appeared to be chiefly of pi 7 'rivA pi
5.8. Authentic markers of these same pi values were glyceralde-
hyde - 3P dehydrogenase (Mr 34,000) and actin (Mr 42,000)
respectively. Both this enzyme and this contractile protein are

abundant in muscle and their mRNA would be expected to exhibit

2+ 2+
good translatability . It is noteworthy that both Cd and Pb ,

singly and combined, increase the translatable mRNA. Some tissue

2+ 2+

responses to Cd combined with Pb have been reported to show

converse effects to those when each substance Is administered

separately.

IV, Fish Enzymes
Animals

Goldfish, Carasslus auratus , were exposed to dechlorinated

2+
acidified tap water (pH 5.0) containing 5, 100 or 500 pg Pb /I as

Pb acetate in a static bioassay for periods of 3, 6 or 20 weeks.

Muscle, liver, kidney and gonads were removed and assayed for 5

enzymes. The muscle and liver were tested for the activity of 2

glycolytic enzymes : glyceraldehyde-3-P dehydrogenase and lactic

dehydrogenase. The liver, kidney and gonads were tested for the

activity of 3 membrane marker enzymes; alanine aminopeptidase,

Y - glutamyltranspeptidase, and alkaline phosphatase. These were

assayed as described previously.

- 377 -

TABLE 1

Lactate Dehydrogenase Activity
(ymoles/mg/hr ± SD)

Dose Time Muscle Liver

(ug/1) (weeks) Control Lead Control Lead

500 3 0.98+0.10 1 . 00+0 . 1 . 82 +0 . 6 1.60 +0 . 1 1

6 1.00+0.08 l.Il±0.08 0.9110.02 1.92 ±0.2o

LOG 3 0.9610J.8 0.94+0.0 6 0.80+0.06 1.37 10.15

6 0.8110.1 5 0.86+0.15 0.81+0.^ 1.52 +0.21

5 20 1.041009 1.15+0.05 0.8510.05 0.9910.01

MeantSD for 2 experiments, each of 3-4 fish.

- 378 -

TABLE 2

Glyceraldehyde-3P Dehydrogenase Activity
(pmoles/mg/hr ± SD)

Dose Time Muscle Liver

(ug/l) (weeks) Control Lead Control Lead

SJDG 3 0.19+0.01 O.Il +O.Ol 0.21±0.02 0.22±O.Ol

6 0.18+0.02 O.lU O.Ol 0.19+0.01 0.22+0.02

im 3 0.1710.01 0.12+0.01 0.21+0.01 0.26+0.03

6 0.1910.02 0.1510.01 0.2310.02 0.20+0.0^

S 20 0.17+0.01 0.1710.01 0.2310.01 0.2310.0:1

MeantSD for 2 experiments, each of 3~^ fish.

- 379 -

TABLE 3

Aminopeptidase Activity
(ymoles/mg/hr ± SD)

Dose Time Liver Kidney Gonads

(pg/l) (weeks) Control Lead Control Lead Control Lead

500 3 A.52±0iJ-2 7.10 ±oM 8.70+0-30 7.10±0.70 8.94+0.42 9.05±0.0 5

I 3.6510.20 7.78i(U5 8.09i0.11 7.67±oj,4 8.36!:0-23 a^9±0.08

100 3 2 . 50±ai 3 . 50*0.05

6 A.47ta22 6.60 10-30 8. 7910,09 8.4010.20 9.0010.30 960rO_10

5 20 4.10+030 4.30+0^0 7.50+0.^5 7.60+0.15 8.70+0s35 8.85+0.7 5

Mean-SD for 2 experiments, each of 3-^ fish.

- 380 -

soo

100

TABLE 4

Y-Glutamyl Transpeptidase Activity
(ymoles/mg/hr ± SD)

Dose Time Liver Kidney ^0"^*^^

(,.g/l) (weeks) Control Lead Control Leac^. Control Lead

1.80±0.0^ 2.30+0.05 2.73+0-23 2.41±0.11 1.1510.0^^ 3.47 ±0.52

1.60-tOi)7 2(^-6 +0.36 2.25^011 2.05t0.D8 1.0 6t0.01 2JA_+0.13

1.82t012 2.40*0-30 2.80+0.^0 2.64^02^1 1.10t0^3 2.00tOp6

1.271020 1.57±0p2 1. 851003 2.00tqil 0.8 It0p6 1.10tOP5

1.60+0.20 I.l5±0i3 2.25+0.25 2.05+0.o5 l.OUO.H 1.21+0.o6

MeantSD for 2 experiments, each of 3"^ fish

- 381 -

RESULTS
At 100 yg Pb '^/l and 500 Mg Pb "*"/l the liver, but not the
muscle, exhibited significant Increases In lactic dehydrogenase
activity after 3 and 6 weeks' exposure. The activity of muscle

glyceraldehyde-3-F dehydrogenase was depressed, especially after

2+
500 ug Pb /I. Similarly the activity of aminopeptidase was

2+
elevated in the liver at these times and levels of Pb exposure.

The kidney and gonads remained of normal aminopeptidase activity.

The liver as well as the gonads exhibited Increase y-glutamyl-

tanspeptidase activity, particularly at the higher Pb level.

Alkaline phosphatase activity was elevated only In the gonads and

not in the other tissues studied, with exposure to 100 yg Pb /I.

2+
The results from 20 weeks' exposure to 5 pg Pb /I water

contrast with the result just described, since no changes In
enzyme activity were detected.

V. SUMMARY

The muscle and kidney of rats exposed acutely or
chronically to Pb or Cd exhibited increases in mRNA
translation. The kidney also showed changes in certain
brush border enzyme activities. These were not correl-
ated with Pb levels in the tissue and were not detected
by urine tests. Fish muscle and liver obtained after Pb
exposure also showed changed enzyme activitiesi suggest-
ing a bioassay.

- 382 -

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27. Johnson, A. J. , Kline» D.L. and Alkjaersig. N. (1969). Assay

methods for plasmin, plasminogen and urokinase in purified
systems. Thromb. Diath. Haemorrh. 21 , 259-272.

28. Maciag, T., Mochan, B., Pye, E.K. and Iyengar, M.R. (1977). IN

Thrcmbosis and Urokinase (R. Paolettl and S. Sherry, eds.) pp.
103-113. Academic Press. New York.

29. Shelton, K.R. and Egle, P.M. (1982). The proteins of lead induced

intranuclear inclusion bodies. J. Biol. Chem. 257, 11802-11807

30. Sims, N.M., Kelley, K.L. , Dayer. J.-M. and Krane, S.M. (1981).

Calcitonin stimulates amino acid incorporation into plasminogen
activator by cultured renal tubular cells. FEBS Lettr. 132,
17^-178.

Acknowledgement

This research was generously supported "by a grant from
the Ontario Ministry of the Environment.

- 385 -

REMOVAL OF HAZARDOUS COHTAMIMANTS IN THE
HAMILTON HATER POLLUTION CONTROL PLANT

G. Zukovs , R.J. Rush, M. Gamble
CANVIRO CONSULTANTS LTD.

Abstract

An assessment Is presented of the Incidence and removability of
selected hazardous organic and Inorganic contaminants (HCs) at the Hamilton
WPCP. The principal study objectives were to evaluate the annual HC loadings
entering and being discharged from the WPCP and to determine the present pro-
cess efficiency for and factors Influencing HC removal. HC monitoring in-
cluded the solid and liquid phase concentrations of total PCBs and other sel-
ected chlorinated organlcs, PAHs and heavy metals. As well, extensive moni-
toring of conventional parameters was conducted in order to characterize both
unit process (e.g. primary treatment) and overall plant performance.

Results indicate a high degree of overall removal (>97% for WPCP as
a whole) for the PAHs. Total PCBs were similarly well removed, averaging
90%. Both lindane and pentachlorophenol were removed to a lesser extent (70%
to 63%) and with considerably less consistency. Heavy metals removals were
generally in excess of 80%.

Results further Indicated that as an overall average of all the HCs
monitored, approximately 20% of the loadings originated from the in-plant re-
turn stream. This varied considerably between specific contaminants being
less than 10% for some PAHs. 52% for pyrene, 12% for nickel and 20-25% for
the other metals.

- 386 -

INTRODUCTION

Background and Relevance

In the modern Industrial society, hazardous contaminants (HCs),
both trace metals and organic compounds are discharged Into public sewage
systems. Until very recently, relatively little was known about tN identi-
ties and quantities of contaminants entering wastewater treatment plants and
little information was available on the factors which would influence their
treatability and ultimate fate.

Recently published reports include four dealing with large scale
field surveys and two based on studies of pilot-scale wastewater treatment
facilities:

1. Survey of 40 Publicly Owned Treatment Works (POTW) by U.S. EPA
Effluent Guidelines Division (EPA 1982a).

2. A 30 day study at a POTW by U.S. Effluent Guidelines Division (EPA

1982b).

3. Survey of 25 POTW by U.S. EPA Municipal Environmental Research

Laboratory (MERL) (Cohen et al 1981).

4. The 5 plant study by the Chemical Manufacturers Association (CMA)
and the U.S. EPA (CMA/EPA 1982).

5. U.S. EPA MERL pilot plant studies of semi-volatile organic com-
pounds (Petrasek et al 1980 and 1981).

6. Pilot plant study of a group of volatile and semi-volatile organic
compounds by van Rensburgh et al (1980) of the National Institute
for Water Research in South Africa.

Results from the EPA surveys and pilot plant studies Indicated that
the wastewater treatment processes studied were stable over a wide range of
operating conditions and were generally effective In removing toxic substan-
ces, as shown in Table 1.

However, the EPA 40 PO™ study (EPA 1982a) showed that Individual
inorganics; e.g. As. Cd, Cu. Hg and Pb, and organics; e.g. polynuclear aroma-
tic hydrocarbons (PAHs) and pesticides pass through a number of treatment
plants in amounts and with frequency to be probable cause for concern.

- 387 -

TABLE 1. EXAMPLES OF TOXICS REMOVAL EFFICIENCY IN MUNICIPALS WPCP's

STUDY

REFERENCE

REMOVAL EFFICIENCY

40 POTW Survey

25 POTW Survey
Pilot Plant Study

EPA (1982a)

Cohen (1981)
Petrasek (1981)

For half of the plants studied:
70% for metals
82% for volatile organlcs
65% for base-neutral organlcs

>80% for many organlcs

>90% for the semi -vol atlles studied

As influent concentrations of many conventional and priority pollu-
tants increased, effluent concentrations also increased. This Implies that
the removal rates for the priority pollutants were relatively constant and
that a fixed percentage of the loading of these pollutants was removed by
secondary treatment.

In general, the higher the industrial contribution to a POTW, the
higher the concentration of priority pollutants in the POTW Influents. Heavy
rainfall increased metallic priority pollutant mass loading at POTWs while
the mass loading of both metallic and organic priority pollutants In POTW in-
fluents was higher on weekdays than on weekends.

Some pollutants not detected in POTW Influents were regularly mea-
sured at high levels in the corresponding sludge streams; e.g. PAHs and
phthalates which were concentrated to the greatest degree In sludges. In
this regard, the survey data [EPA (1982a) and Cohen (1981)] support the find-
ings of Petrasek (1981), who has suggested that sludges, particularly primary
sludges are likely to act as a sink for these compounds because of a blocon-

centratlon effect.

In the South African study (van Rensburg et al 1980), van Rensburg
reported 90 percent effectiveness In the removal of toxic organics even under
the pressure of shock loads of these chemicals, and he observed a severe
build-up of some compounds in the recycled sludge. Both of these observa-
tions are in agreement with those reported by the U.S. EPA in their surveys
and pilot-scale studies (EPA 1982a, Cohen 1981 and Petrasek 1981).

Trace organics monitoring studies at several Canadian sewage treat-
ment plants have been carried out (e.g. EPS 1980 and MOE 1980) and continuous
flow fate studies of trace organics in a pilot plant at the Wastewater Tech-
nology Centre (WTC) in Burlington, Ontario are on-going. However, currently.

- 388 -

In Canada the data base for HCs occurrence and removal at WPCP*s 1s ^ery
limited {CANVIRO 1983). Therefore. In August 1982 a study was Initiated to
provide an accurate estimate of the annual loading of HCs entering and being
discharged from the Hamilton WPCP.

STUDY OBJECTIVES

In previous toxics studies of this nature, numerous analytical
problems have been reported, thus, the study was divided Into two phases.

Phase 1 objectives were to Identify the most significant HCs pre-
sent in the WPCP, to establish a list of HCs for the study's monitoring pro-
gram and to optimize site sampling techniques and analytical methods. Phase
2 objectives were to estimate present treatment process efficiency and to de-
termine the factors influencing contaminant removal.

In the fall of 1983. while the study was still In progress. CANVIRO
prepared a preliminary paper on the results of Phase 1 (Rush and Taylor,
1983). It presented a summary of the experimental procedures adopted for the
study, some analytical QA/QC results and preliminary treatment plant effi-
ciency data.

In contrast, this paper discusses the results of the second phase
of the study and presents an assessment of the Incidence, and treatability of
selected hazardous organic and inorganic contaminants in the Hamilton WPCP.
The factors affecting contaminant removals are also discussed herein.

STUDY PROCEDURES

Saiipling Methods

Twenty-four hour composite samples were collected during three sep-
arate periods in Phase 2, representing three seasonal periods, as shown in

Table 2.

TABLE 2. PHASE 2 SAMPLING PROGRAM SCHEDULE

PHASE 2

SAMPLING PERIOD

NUMBER OF SAMPLING DAYS

Winter
Spring
Summer

December 1982 - January 1983
April and May 1983
June - August 1983

6
4
4

- 389 -

The sampling locations In the WPCP are shown In Figure 1, and the
following samples were collected:

1. Combined Influent; I.e. raw sewage combined with the In-plant re-
turn stream.

2. Effluent.

3. Waste activated sludge (WAS).

The In-plant return stream was also sampled on an intermittent
basis. The WPCP operation was observed closely during the sampling periods
and the process data together with observations of any WPCP operational up-
sets were recorded for correlation with the analytical results for the para-
meters being measured.

This study was unique, In that the solid and liquid fractions of
the samples were examined separately. Therefore, the samples were centri-
fuged to separate the two fractions ready for analysis. However, the suspen-
ded solids concentration of the effluent was extremely low, and in order to
avoid centrifuging very large volumes of effluent, WAS solids were used in
place of effluent solids (assuming the two to be equivalent).

The contaminants monitored during the study Included total PCBs,
selected pesticides, six PAHs, two polynuclear heterocyclic organics, trace
metals and a group of conventional parameters. Table 3 presents a complete
listing of the specific compounds monitored.

STUDY RESULTS

Treatment Plant Operation During the Study Period

In order to provide background on the treatment plant operation and
performance during the study period, all the routine monitoring records from
the facility were reviewed and summarized. Analysis of these records for
variability and comparison to plant design parameters allowed the characteri-
zation of both unit process and overall plant performance.

A summary log of major operational events was developed to monitor
the effect of equipment failures, equipment shutdowns and other events on the
WPCP's efficiency. A relatively small number of operational events were
identified during the study period, none of which appeared to significantly
effect the plant HC removal efficiency.

CHLORINE
CHAMBER

DETRITORS

1 II / R

W^=^

TOTAL
INFLUENT

PUMP IL

HOUSE

I r F

^^ — ^ /> 'A

IN PLANT />^ IT

RETURN ^y |

AERATION

TANKS

CI./

riTTT

I r R

RETURN
ACTIVATED
SLUDGE
(RAS)

<g)

WASTE ACTIVATED
SLUDGE (WAS)

r " 'EFFLUENT SEWER , * _ ."".J

FINAL
EFFLUENT

ASH QUENCH WATER

OTHER INPLANT WATERS

FILTRATE

SUPERNATANT

L E G E N D

0-SAMPLING LOCATION

ASH TO
LANDFILL

INCINERATOR

SLUDGE FILTERS

COMBINED PRIMARY

AND WAS FROM
PRIMARY CLARIFIERS

FIGURE 1-WPCP PROCESS FLOW SCHEMATIC AND SAMPLING LOCATIONS

- 391 -

TABLE 3. HAZARDOUS CONTAMINANTS MONITORED DURING THE STUDY

COMPOUND

CLASSIFICATION

Acenaphthylene

Benzo{a)pyrene

Fluorene

Fluoranthene

Naphthalene

Pyrene

Polynuclear Aromatic

Hydrocarbons

(PAHs)

Carbazole
Dibenzofuran

Polynuclear Heterocyclic
Compounds

G-BHC (Lindane)
Total PCBs
Pentachlorophenol
Other Pesticides *

PCBs and Pesticides

Aluminum

Arsenic

Cadmium

Calcium

Chromium

Copper

Iron

Lead

Magnesium

Mercury

Nickel

Selenium

Zinc

Trace Metals

* other Pesticides:

Aldrin

A-BHC

B-BHC

A-Chlordane

G-Chlordane

Dieldrin

DMDT Methoxychlor

Endosulfan I

Endosulfan II

Endrin

Endosulfan Sulphate

Heptachlorepoxide

Heptachlor

Mi rex

Oxychlorodane

OP-ODT

PP-DDD

PP-DDE

PP-ODT

Hexachlorobenzene
2 ,3 ,4-Trichl orophenol
2.3,4,5-Tetrachlorophenol
2 ,3 ,5 ,6-Tetrachl orophenol
2, 4, 5-Tr1chl orophenol
2 ,4 ,6-Tr1chl orophenol

- :^92 -

Three separate sets of performance data were recorded at the WPCP
during the study period. These Included average monthly routine monitoring
data, average conditions one week prior to sampling and actual sampling day
data. In general, the sampling day results were in good agreement wit.i aver-
age weekly and monthly monitoring values. For example, less than a 'dO% var-
iation was recorded in influent suspended solids and BOD5 co.^critrations,
while effluent levels for the same parameters were within 10% for all three
data sets.

Analysis of the combined operational and performance monitoring
data verified that the Hamilton WPCP was functioning well and within typical
(for the Hamilton WPCP) operating ranges on the sampling days of the study.
Thus, the data collected is believed to provide "typical" results from that
plant.

OVERALL REMOVALS OF HCs

Because of the volume of results generated in this study, space
does not permit the inclusion of all the influent and effluent concentra-
tions, masses and percent removals summary tables for all the individual con-
taminants on each sampling day in this paper. A summary of the average and
range of influent and effluent concentrations of all contaminants monitored
in this study is presented in Table 4. The removal efficiencies for each
contaminant during the seasonal periods of the study are discussed below.

Polynuclear Arowatlc Hydrocarbons (PAHs) Rewovals

Table 5 provides a summary of the removal efficiencies for PAHs in
both the solid and liquid fractions of the wastewater as well as indicating
the overal 1 percent removal s . For convenience . the two heterocycl ic com-
pounds, carb-izole and dibenzofuran have been grouped with the PAHs for dis-
cussion throughout the remainder of this paper.

The overall removals of the individual PAHs ranged from 94X for
naphthalene to essentially lOO'^S for acenaphthylene. The overall percent re-
moval fc- the combined group of PAHs was 97'E. Removals within the liquid and
solids fractions were in the same range, being divided approximately equally
between the two fractions for all the compounds except for naphthalene which
had a slightly lower average removal (89%) in the liquid fraction. This was

- ?>9'S -

TABLE 4. AVERAGE CONCENTRATIONS OF HCs ENTERING AND BEING
DISCHARGED FROM THE HAMILTON WPCP

CONTAMINANT

INFLUENT

EFFLUENT

AVERAGE
CONCENTRATION
(ug/L)

CONCENTRATION
RANGE
(ug/L)

AVERAGE
CONCENTRATION
(ug/L)

CONCENTRATION
RANGE
(ug/L)

Naphthalene
Acenaphthylene
Dibenzofuran
Fluorene

13.43

5.76

10.92

14.46

0.83-119.1
0.54-38.45
0.99-115.2
1.42-154.7

0.28
0.04
0.12
0.19

0.0-0.68
0.0-0.54
0.0-0.99
0.0-2.05

Fluoranthene

38.70

3.18-418.8

0.61

0.05-3.13

Carbazole

21.62

7.68-76.03

0.41

0.0-1.07

Pyrene

35.25

3.01-379.2

0.80

0.0-4.96

Benzo{a)pyrene

41.10

11.87-129.0

0.62

0.0-2.70

G-BHC (Lindane)

0.09

0.0-0.29

0.03

0.0-0.16

Total PCBs

0.13

0.0081-0.40

0.03

0.0-0.30

Pentachlorophenol
Other Pesticides*

0.23
<0.005

0.0-1.01
0.0-0.005

0.10
<0.001

0.0-0.29

0.0-0.001

Iron

6.931

1.846-16,191

488.6

271-1.719

A1 umi num

2.220

434-6.413

363

143.9-814.2

Arsenic

Calcium

1.88

71,329

0.0-5.81
67.785-94.713

0.12
61.632

0.0-0.52
50,508-91,148

Cadmium

0.99

0.09-1.86

0.08

0.03-0.36

Chromium

206.2

71.0-559.8

18.24

0.0-67.11

Copper

131.4

24.9-243.1

18.38

3.26-42.69

Mercury

0.26

0.0-0.36

0.03

0.0-0.11

Magnesium
Nickel

29.019
91.27

17.029-161,444
56.3-149.6

35.847
37.13

13.000-160.000
0.97-64.6

Lead

94.15

8.81-147.9

6.19

0.0-16.52

Selenium

1.45

0.0-4.77

8.82

0.0-120.0

Zinc

3,283

254-40.269

91.34

36.80-147.6

Other Pesticides Include:

Al drin

A-BHC

B-BHC

A-Chlordane

G-Chlordane

Die! drin

DMDT Methoxychlor

Endosulfan I

Endosulfan II

Endrin

Endosulfan Sulphate

Heptacniorepoxide

Heptachlor

Mi rex

Oxychlorodane

OP-DDT

PP-ODD

PP-ODE

PP-ODT

Hexachlorobenzene

2,3,4-Trichlorophenol

2.3,4,5-Tetrachlorophenol

2,3.5,6-Tetrachlorophenol

2,4.5-Trichlorophenol

2,4,6-Trichlorophenol

TABLE S, OVERALL Rt«0.*L EFFKiENCr FOR PAilS

NAPHTh*; rK

ftiEh

:PHTrifLENt

DlbEhZOFDHAM

FLUOREHE

FLIIORANTHENE

CARBAZOLE

PYRENE

BE:NZOIa)PrRENE

t'Lr':LHT RtM.'lVAL

^ikat,', REMLlVAl

PEfiCtNT REMOVAL

PERCENT REMOVAL

b FKAC
99727

L fkAC

luTii:.
"^9722"

L FhAC
100.00

TOTAL
100.00

I EKAC

100. on

L FRAC

TOTAl

S FRAC

L FRAC

TOTAl

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

Dft: 06-07

100.00

99.54

100-00

99.71

88.64

60.00

83,62

94.19

100.00

99.69

89.02

50.00

63.77

97.29

93.75

96.61

Iiet ?:j-21

9'*.9b

INV

UlO OU

lllO.OO

liiO.OO

100. 00

9b. 15

99.14

99.96

96,63

98.67

99,21

99.41

99.25

96.67

100.00

98.90

98.69

99.23

98.69

99.20

-Jan iJ',-[iL'ts

99. J9

9J .'i>

99. 6 J

IJO.Ou

99.91

99.67

100.00

99. B2

99.70

100.00

99.86

93.17

83.33

91.51

89.39

100.00

96.12

86.88

75.00

86.65

97.64

100.00

98.15

Jan 11-1?

1 ao , 00

IOC. GO

IOC. CO

ILCOu

100.00

96.02

lOO.OU

97.56

100.00

100,00

100.00

9S.34

lOU.OO

95.91

100.00

94.60

100.00

95.38

94.71

100.00

96.67

Jan 18-19

9&.70

9tS,7D

99. 3i

100.00

99.76

99. 2E,

100.00

99.65

99.63

100,00

99,77

96.57

100.00

97.04

92.49

100.00

96.86

95.64

100.00

96.07

98.54

100.00

99.07

Jan ?5-26

WINTER
AVERAGE

100.00
99.63

INV

100. OU

100.00

99.63

100.00

99.8c)

99.85

lOO.OO

99.92

97.66

100.00

97.86

98.10

100.00

98.59

97.11

100.00

97.33

100.00

100. 00

100.00

99.44
idu.Dij

99.84

" 97 ."tV

100.00

99.94

99 13

99.69

99.34

99.78

99.47

99.66

95.08

88.79

94.20

95.14

100.00

97.86

93.97

87.37

93.02

97.90

98.75

98.26

Apr 13-14

100.00

IQO.OO

100. uo

99.49

97.29

100.00

98.52

98.71

100.00

99.21

95.85

100.00

96.49

93.38

100.00

96.16

94.05

100.00

94.90

97.18

100.00

97.86

Apr 21-22

100.00

100. 00

100.00

too. 00

100.00

98.02

100.00

98.32

98.29

100.00

99.17

97.63

100.00

99.97

99.09

100.00

99.24

Apr Z6-27

100.00

94 44

94,79

100.00

50.00

100. 00

50.00

66.09

97.97

100.00

96.30

97.75

97.06

97.35

97.19

100.00

97.65

99.09

100.00

99.32

Hay 02-03

100.00

50.00

6i.22

99.25
INV

100.00

100 . 00

100.00

93.40

100.00

94.21

91.98

95.00

93.69

91.98

100.00

93.60

96.26

100.00

97.05

SPk 1 HG

AVERAGE

100. OU

Q6.11

90. UO
TV

100.00

99.83

99.10

B7.50

87.13

99.68

87.50

91.82

96.31

100.00

96.83

95.35

98.02

96.60

95.21

100.00

96.03

97.91

100.00

98.37

June 22-23

96.60

INV

100. OU

100,00

100.00

99.69

100.00

99.79

99.02

100.00

99.13

97.96

100.00

98.87

98.31

100.00

96.56

98.68

100.00

98.92

July 11-12

100.00

63.33

86.09

100.00

97,97

100.00

98.58

95.01

100.00

99.21

96.96

100.00

97.93

97.91

100.00

96.71

tug 02-03

97.23

95.93

96.t)9

100.00

too. 00

100.00

9^.47

100.00

98.01

96.91

100.00

99.36

97.23

100.00

97.88

99.78

100.00

99.67

Aug OB' 09

96.64

100.00

99.87

93.08

100.00

99.37

100. OU

100.00

96.70

100.00

97.73

93.47

100.00

97.95

96.31

100.00

97.00

99.02

100.00

99.39

SUHHCR

Average

97.34

93.06

93.62

96.54

100.00

99.79

JOQ.OO

100.00

99.92

100.00

99.95

97.79

100.00

98.36

95.84

100.00

98.85

gv?©

100.00

97.85

98.85

100.00

99.22

OVERALL

A/LriAGE

99.04

H9.09

94.84

99.08

100.00

99.88

99.39

96.30

96.04

99.79

96.20

97.50

96.2

95.20

96.14

95.40

99.43

97.98

95.. "'5

94.59

95.26

96.17

99.52

96.58

MAI 1 HUM

100.00

100,00

100.00

99.2

100.00

99.25

100.00

98.59

100.00

98.69

100.00

HINIHUH

96.60

60.00

65.22

93.08

ioo.no

99.37

96.02

50.00

98.71

50,00

68.09

88.54

50.00

83.62

89.39

95.00

93.69

86.88

50.00

83.77

94.71

93.75

96.61

Notc&j^ INV - Invalid Result

TV - Only Trace Values detected

- ^9S -

partially due to the extremely low concentrations of naphthalene in the
liquid fraction during the winter sampling period. It should be noted that
three isolated results of uncharacteristically low PAH removal efficiencies
were recorded in Table 5 (50% for dibenzofuran on April 26-27, 65.2% for
naphthalene on May 2-3, and 68.1% for fluorene on April 26-27). The low re-
movals were all observed on sampling days when influent loadings of the spe-
cific contaminants were very low and corresponding effluent levels were at or
nedr average values. Therefore, despite the typically low concentrations
present in the effluent, the associated removal efficiencies calculated for
those periods were extremely low.

The treatment plant efficiency for acenaphthylene and benzo(a)-
pyrene removal was extremely good. Efficiencies for the two contaminants
were In the 98-100% range and the effluent streams contained only traces of
either compound (e.g. <1.0 ug/L).

Fl uoranthene, fl uorene and pyrene 1 nf 1 uent level s were somewhat
higher than the to 12 ug/L range found In the EPA 30 day study (U.S. EPA
1982). Effluent values varied from almost non-detectable levels of fluoran-
thene to a 5.0 ug/L concentration of pyrene resulting from the excessive
loading of December 20-21. The overall removal efficiencies of 95-96% re-
flected the systems ability to consistently remove fluoranthene, fluorene and
pyrene at the Influent concentrations encountered (I.e. 1.5 ug/L - 420 ug/L).

The system generally responded well to large doses of trace orga-
nlcs, actually attaining higher than average removal rates for the increased
loads of the December 20-21 period.

PCBs and Pesticides

Table 6 provides a summary of removal efficiencies for three com-
pounds; lindane, total PCBs and pentachlorophenol (all other pesticides were
below detection limits in most samples).

Removals of lindane ranged from a low of 16.7% to 100% and did not
exhibit any distinct trends. The large variance In treatment efficiency was
probably related to the minute quantities of lindane which were detected.
Influent concentrations in both the solid and liquid phase were, all below
1.0 ug/L. Consequently, effluent values were usually below the detection
limit, making the value of the removal data questionable.

TABLE 6. OvlRALL RtMOVAL EFFICItNCY FOR PCBs AND PESTICIDES

G-BHC
PERCENT REMOVAL

TOTAL PCB
PERCENT REMOVAL

PENTACHLOROPHENOL
PERCENT REMOVAL

S. FRAC.

TV
TV
TV
TV
TV
TV

L. FRAC.

INV
100.00
43.19
64.03
86.86
91.67

TOTAL

S. FRAC.

L. FRAC.

TOTAL

S. FRAC.

L. FRAC.

TOTAL

Dec 06-07
Dec 20-21
Jan 05-06/83
Jan 11-12
Jan 18-19
Jan 25-26

INV
100.00
43.19
64.03
86.86
91.67

81.21
71.43
100.00
94.17
91.31
97.78

INV
76.19
76.47

TV
100.00

INV
74.26
78.74
94.76
91.31
99.48

100.00

95.76

100.00

TV
100.00

TV
11.11
30.43
11.11

ND
100.00
95.76
29.82
30.43
45.68

WINTER
AVERAGE

INV

77.15

89.32

84.22

87.59

98.94

17.55

60.34

Apr 13-14
Apr 21-22
Apr 26-27
May 02-03

100.00
INV

100.00
NO

100.00

INV
42.11
ND

100.00

INV
42.21
ND

100.00
INV
99.35
INV

TV
73.33
100.00

INV

100.00
INV
99.62

INV

INV
86.73
ND
ND

44.12

TV
ND
ND

INV
86.73
ND
ND

SPRING
AVERAGE

100.00

71.10

99.68

86.67

99.81

86.73

44.12

86.73

June 22-23
July 11-12
Aug 02-03
Aug 06-09

100.00

16.67
85.71

ND
16.67
86.12

TV
81.99
93.37
93.18

100.00
42.86
63.64

100.00

100.00
71.41
84.51
94.29

TV
97.54
91.26

INV

65.88

20.00

80.00

INV

65.88

32.05

80.15

INV

SUMMER
AVERAGE

OVERALL
AVERAGE
MAXIMUM
MINIMUM

100.00

51.19

51.40

89.51

76.63

87.55

94.40

55.29

59.36

100.00
100.00
100.00

70.02

100.00

16.67

70.08

100.00

16.67

91.25

100.00

71.43

81.39

100.00

42.86

89.80

100.00

71.41

95.90

100.00

86.73

37.52
80.00
11.11

62.94

100.00

29.82

Notes: ND = Not Detected
INV = Invalid Result
TV = Trace Value

- 397 -

Shannon (1976) examined PCB concentrations in 33 wastewater treat-
ment plants 1n Ontario and found that Influent loadings ranged from the de-
tection limit (0.01 ug/L) to 1.8 ug/L. Comparison of the influent PCB values
in Table 4 shows that the PCB input into the Hamilton WPCP was in the range
reported by Shannon (1976), with an overall average of 0.13 ug/L of PCB being
present In the Hamilton influent during this study period.

The overall removal efficiency of polychlorinated bi phenyls was
good (90%) with an overall average concentration In the effluent of 0.03 ug/L
during the study.

Pentachlorophenol (PCP) differed from the PCBs in that it was pre-
sent mainly in the liquid fraction of the wastewater. Influent concentra-
tions ranged from ug/L to 1.01 ug/L with the highest levels being recorded
during the winter. Although an overall removal efficiency of 63% for penta-
chlorophenol was determined from the testing, on days when the Influent con-
tained more than trace amounts of the compound, removal was over 80%. This
meant that effluent levels of pentachlorophenol were quite low (average = 0.1
ug/L) and were commonly at or below the detection limit (0.001 ug/L).

Trace Metals

Table 7 provides a summary of the removal efficiencies for metals
in both the liquid and solid fractions of the wastewater as well as indicat-
ing the overall percent removals. Results were not obtained for trace metal
concentrations on the second sampling day (December 20-21), and thus, the
metals analysis is based on thirteen samples rather than fourteen.

Removals of trace metals in the WPCP correlated well to influent
concentrations as periods of high removal efficiency corresponded directly to
periods of high metal loadings. Brown (1973) found the same trend in his
study of the efficiency of municipal sewage treatment plants, in which he
examined the removai of five metals (Cu, Cr, Zn, Pb, Cd).

Trace metal removals were better accomplished In the solid fraction
of the wastewater than in the liquid fraction. However, there was some vari-
ance in this trend due to fluctuations in the liquid/solid composition of the
influent wastewater. In addition to composition variation, deviations in the

lAfiLE I. OVERALL REMOVAL EFFICIENCY FOR TRACE ftTALS

IRON
PERCENT RDluV.Al.

AlllfllNUM
PERCENT REMOVAL

ARSENIC
PERCENT REMOVAL

CADMIUM
PERCENT REMOVAL

CHROMIUM
PERCENT REMOVAL

COPPER
PERCENT REMOVAL

S FRAC

L FRAC

TOTAL

ND
93.42
93.02
92.69
97.02

94.04

S FRAC

38.88
ND
94.59
94.14
87.26
99.85

92.94

L FRAC

48.84
ND
95.36
60.78
45.95
51.35

60.46

TOTAL

81.91
ND
95.09
87.99
77.13
92.64

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

S FKAC

L FRAC

TOTAL

Dec 06-07
Dec 20-21
Jan 05-06/8J
Jan 11-12
Jan 18-19
Jan 25-26

WINTER
AVERAGE

Apr T3-14
Apr 21-22
Apr 26-2/
May 02-03

SPRING
AVERAGE

June 22-23
July 11-12
Aug 02-03
Aug 08-09

surtiER

AVERAGE

ND
94.34
93.88
92.55
97.4/

94.56

97.35
97.99
97.80
93.88

96.75

ND
92.12
96.02
93.65

NIJ
90.00
89.41
93.20
92.6/

91.32

INV

70.00

96.52

100.00

88.84

83.36
ND

92.52
ND
ND
ND

TV
ND
TV
ND
ND
ND

83.36

ND
92.52

77.70
ND
93.54
87.94
89.57
94.13

TV
ND
TV
TV
TV
TV

77.70
ND
93.54
87.94
89.57
94.13

82.89
ND
92.16
92.95
86.93
95.61

90.00

ND
100.00

83.33
100.00
100.00

85.29
ND

93.37
90.47
89.95
96.47

88.85
ND
94.23
92.91
90.00
96.27

0.00

100.00

25.00

83.33

100.00

83.72
ND

95.06
77.75
88.50
96.70

86.95

87.94

INV

87.94

88.58

INV

88.58

90.11

94.67

91.11

92.65

61.67

88.35

INV
91.16
97.53
94.53

94.41

94.42
96.89
97.90
92.18

52.63
INV
8.70

96.81

87.08

INV
87.85
95.57

94.13
95.14
99.64
91.04

TV
TV
TV
TV

94.13
95.14
99.64
91.04

93.83
97.29
98.26
90.94

TV
TV
TV
TV

93.83
97.29
98.26
90.94

93.77
96.31
96.91
93.53

100.00

TV
100.00
100.00

94.67
96.31
97.53
95.15

95.59
97.19
97.80
91.09

INV

87.50

100.00

INV

94.41

93.35

95.34

52.71

90.17

94.99

INV

94.99

95.08

INV

95.08

95.13

100.00

95.92

95.42

93.75

93.88

ND
INV
95.24
88.33

ND
INV
95.77
92.65

"95.75^
92.34
96.33
94.60

26.67

INV

26.47

26.92

83.35

INV

71.43

77.42

98.96

4.24

94.19

97.05

TV
TV
TV
TV

98.96

4.24
94.19
97.05

93.23

92.02
95.57
93.37

TV
TV
TV
TV

93.23
92.02
95.57
93.37

92.39
94.48
97.97
94.01

100.00
76.92
71.43
91.67

93.46
78.26
88.01
93.24

94.00
91.72
96.58
95.46

100.00
INV
50.00
33.33

94.93

INV

86.86

84.34

93.93

91.79

94.21

94.63

26.69

77.40

73.61

INV

73.61

93.55

INV

93.55

94.71

85.00

88.24

94.44

61.11

88.71

OVERALL
AVERAGE
MAXIMUM
MINIMUM

95.19
97.99
92.12

90.60

100.00

70.00

94.20
97.53
91.16

94.20
99.85
87.26

49.13

96.81

8.70

85.22

95.57
77.13

85.23

99.64

4.24

INV
INV
INV

85.23

99.64

4.24

92.11
98.26
77.70

INV

INV
INV

92.11
98.26
77.70

93.07
97.97
82.89

92.78

100.00

71.43

91.71
97.53
76.26

94.05
97.80
88.85

67.92

100.00
0.00

89.56
96.70
77.75

Notes:

ND = Not Detected
INV = Invalid Result
TV = Trace Value

TABLE 7 (CONT'D). OVERALL REMOVAL EFFICIENCY FOR TRACE METALS

['.lRCURY
PERCENT REMOVAL

NICKEL
PERCENT REMOVAL

LEAD
PERCENT REMOVAL

SELENIUM
PERCENT REMOVAL

ZINC
PERCENT REMOVAL

S FRAC
94.98

L FRAC
100.00

TOTAL

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

S FRAC

L FRAC

TOTAL

Dec 06-07

95.60

88.33

20.00

50.70

88.83

50.52

84.53

75.00

82.85

Dec 20-21

Jan 05-06/83

95.76

82.35

92.15

94.84

40.00

63.66

94.75

93.28

93.64

71.43

88.98

Jan 11-12

96.02

40.00

74.34

94.78

91.76

68.75

85.28

Jan 18-19

91.24

14.29

47.36

95.26

INV

88.08

66.67

80.87

Jan 25-26

96.21

100.00

96.77

97.19

16.67

64.90

96.89

100.00

97.55

95.18

44.44

84.77

WINTER
AVERAGE

95.65

94.12

94.84

93.52

26.19

60.19

94.10

100.00

93.98

71.90

90.64

65.26

84.55

Apr 13-14

95.71

75.00

93.09

92.81

20.00

45.93

95.61

75.00

89.17

90.11

93.75

INV

Apr 21-22

96.94

50.00

89.88

98.27

97.82

INV

93.33

INV

97.19

99.65

99.63

Apr 26-27

98.36

42.86

90.49

98.00

20.00

61.47

98.26

99.25

97.54^

68.75

88.98

May 02-03

SPRING

AVERAGE

90.23

33.33

85.15

93.35

60.00

77.23

95.10

84.20

93.24

69.57

86.38

95.31

50.30

89.65

95.61

33.33

70.73

96.70

75.00

94.18

93.22

91.19

95.43

79.32

91.66

June 22-23

97.96

0.00

65.10

97.77

100.00

98.01

92.06

95.70

15.38

66.92

July 11-12

93.08

11.11

67.30

95.96

14.29

19.57

96.84

INV

95.53

39.13

44.50

Aug 02-03

95.13

0.00

70.68

96.29

INV

96.25

90.86

96.63

58.33

77.81

Aug 08-09

50.00

95.48

33.33

70.66

96.83

100.00

97.31

84.31

93.94

44.44

74.42

SUMMER
AVERAGE

94.11

20.37

62.66

96.47

15.87

51.78

96.92

100.00

97.10

89.08

95.45

39.32

65.91

OVERALL

AVERAGE

95.16

54.38

83.11

95.06

25.32

61.60

95.77

93.75

95.17

87.10

INV

85.57

93.59

60.13

80.12

MAXIMUM

98.36

100.00

96.77

98.27

60.00

98.27

98.26

100.00

98.26

99.25

INV

99.25

97.54

99.65

99.63

MINIMUM

90.23

0.00

50.00

88.33

0.00

19.57

88.83

75.00

88,83

50.52

INV

50.52

84.53

15.38

44.50

Notes: ND = Not Detected
INV = Invalid Result
TV = Trace Value

- 400 -

degree of metals loading were also common. For example. Influent concentra-
tions of aluminum varied from 434 ug/L on July 11-12 to 6413 ug/L on May 2-3,
while iron levels ranged from 1846 ug/L on July 11-12 to 16,191 ug/L on Janu-
ary 25-26. These observed deviations support Oliver and Cosgrove's (197:,
assertion that If a municipal treatment plant serves a heavily industrialized
area, then trace metal Inputs to the plant will occur In isolate,^ :'^ugs. The
slugs are the result of non-periodic high discharges from various contribut-
ing Industrial sources.

The best metal removal efficiencies were achieved for Iron and lead
(95% average removal). However, due to high Influent concentrations, the
effluent concentrations of iron were also high (488 ug/L average). The pre-
sence of iron contributes to the precipitation of phosphorus within the
plant, and thus, the addition of a phosphorus removal chemical is not prac-
ticed at the Hamilton plant to meet the phosphorus objective of 1 mg/L.

A group of metals including arsenic, aluminum, cadmium, chromium,
copper, mercury, selenium and zinc were all removed effectively by the WPCP
with percent reductions ranging from 80% for zinc to 92% for cadmium. It is
noteworthy that despite the industrial presence In Hamilton, these metals all
had influent and effluent concentrations below the average levels recorded by
Environment Canada in its study on metal sources in municipal wastewaters of
Ontario (Environment Canada 1978).

Suniary of Annual Loadings of HCs

An evaluation of the annual loadings of hazardous contaminants
entering the Hamilton WPCP is given in Table 8. The annual loadings were
developed for each specific contaminant and were calculated from the overall
mean loadings for the sampling days. The annual Influent PAH loadings ranged
from 241 kg for acenaphthylene to 3457 kg for benzo(a)pyrene, while annual
effluent masses ranged from 4.1 kg for acenaphthylene to 84 kg for pyrene.
Acenaphthylene and carbazole were the only PAHs which concentrated to a
greater extent in the liquid fraction of the influent wastewater. However.
acenaphthylene, dibenzofuran, fluorene and naphthalene were all associated
with the liquid fraction of the effluent.

- 401 -

TABLE 8. ESTIMATED ANNUAL LOADINGS OF HAZARDOUS CONTAMINANTS ENTERING
AND BEING DISCHARGED FROM THE HAMILTON WPCP

HAZARDOUS
CONTAMINANT

ANNUAL LUAUING (kg/yr

INFLUENT

EFFLUENT

TRACE ORGANICS

SOLID
FRACTION

LIQUID
FRACTION

TOTAL

SOLID
FRACTION

LIQUID
FRACTION

TOTAL

Naphthalene

1.004

427

1,431

1.1

29.2

30.3

Acenaphthylene
Dibenzofuran

241
606

372
544

613
1,150

0.4
1.4

3,7
11.0

4.1
12.4

Fluorene

887

639

1,526

0.7

18.3

19.0

Fluoranthene

3.325

741

4,066

54.8

11.0

65.7

Carbazole

924

1.383

2,310

36.5

7.3

43.8

Pyrene
Benzo(a)pyrene

Lindane

3,110
3,457
0.0

595
960
7.3

3,705
4,420
7.3

73.0
65.7

0.0

11.0
3.7
3.7

84.0

69.4

3.7

Total PCBs

11.0

3.7

14.7

1.1

2.9

4.0

Pentachlorophenol

1.5

21.9

23.4

0.0

11.0

TRACE METALS

Iron

612,105

130.305

742,410n

26,061

24,966

51,027

Aluminum

151.840

98,915

250,755

8,724

30.587

39,311

Arsenic

219

Cadmium

110

Chromium

15.951

67,160

83,111

1.095

913

2,008

Copper

11.826

2.665

14.491

712

1.314

2,026

Mercury

21,9

7.3

29.2

1.1

2.6

Nickel

5,074

4.782

9.892

237

3.723

3.979

Lead

9,782

803

10.585

475

219

694

Selenium

168

183

Zinc

33,580

312,805

346.385

2.482

7.556

10.038

* These estimates ^re based on average values from 14 sampling days and in-
clude very high values measured on one or two days for most contaminants.
Thus, these annual averages are considered to be high estimates.

The annual PCB loading entering the Hamilton WPCP was 14.7 '<g/yr.
Almost 75% of the influent polychlorinated biphenyls were concentrated in the
solid phase of the wastewater, while on the contrary, 75. of the effluent PCB
mass was concentrated in the liquid phase.

- 402 -

The trace metals could be divided Into two distinct groups In terms
of loading quantity;

i) a group Including Iron, aluminum, chromium, copper, niclce'' , lec
and zinc which were consistently present In large amounts {e.g.
5.000-700,000 kg/yr in the Influent; 600-50,000 kg/yr In the efflu-
ent); and
11) a group Including arsenic, cadmium, mercury and selenium which were
intermittently present In small quantities (e.g. 20-300 kg/yr in
the Influent; 5-50 kg/yr In the effluent).

The trace metals In the first group (e.g. iron, etc) concentrated
to a considerable degree in both the solid and liquid fractions of the waste-
water, while the metals in the second group, with the exception of selenium.
existed mainly in the liquid fraction of both the Influent and effluent.

ASSESSMENT OF FACTORS AFFECTING REMOVALS OF HAZARDOUS CONTAMINANTS

General

The factors which potentially affect the removal of hazardous con-
taminants in municipal WPCP's have been broadly categorized into two groups
for purposes of this study:

i) Seasonal effects and miscellaneous factors including Influent con-
taminant concentrations.
ii) Design and operational characteristics of the WPCP.

In addition, the behaviour of the various contaminants monitored
during this study was expected to differ during treatment due to physical and
chemical characteristics which affect how completely and by what mechanism
they are removed. This was assessed in conjunction with each of the two main
factors investigated.

Seasonal Effects

Although it is generally recognized that temperature fluctuations
influence the removal of BOD5 and ether conventional parameters, the results

- 403 -

of this study provide no definite evidence that would suggest seasonal chan-
ges have any significant effect on the treatability of HCs within the Hamil-
ton WPCP. It was found that generally the removals of all HCs studied were
good (e.g. >90% for organics; 80% for metals) regardless of the season.

However, there was a trend toward higher influent contaminant con-
centrations during the winter period of this study, which resulted In slight-
ly higher winter average effluent concentrations for some contaminants (I.e.
since the percent removal was essentially the same from season to season).

WPCP Design/Operational Factors

Design and operational features of the Hamilton WPCP were assessed
to determine their effects on HC removal during this study. These Included:

1) SRT, HRT and hydraulic loading
1i) Primary versus secondary removals
1i1) In-plant return stream loadings

Effect of SRT. HRT and Hydraulic Loading

A sunmary of the Important factors SRT. HRT and final clarlfler
hydraulic loadings during the seasonal periods of the study Is presented in
Table 9 along with the average and ranges of removals achieved for the PAHs
as a group, total PCBs and a group of seven metals (iron, aluminum, chromium,
copper, nickel, lead and zinc). Other metals were present in very low con-
centrations and were omitted from this assessment, as were lindane and PCP.

During the study the SRT ranged from 2.9 days to 8.8 days with an
overall average of 5.7 days. The winter, spring and summer average SRTs did
not differ greatly, being 5.4, 4.9 and 5.7 days, respectively.

The average HRT in the aeration section of the plant varied some-
what throughout the study depending mainly on the aeration volume in ser-
vice. The winter, spring and summer averages ^ere 2.47, 3.14 and 3.74 hours,
as more aeration volume was utilized during the summer period.

One of the most notable changes in the plant operation during the
study was the balancing of flow between the old section and new section final
clarifiers. As shown in T^rble 9 during the winter and spring periods, the

TABLE 9. SUMflARY OF KEY DESIbM/OPERATlONAL PARAMETERS VERSUS HC REMOVAL

AVLKAOE AERATION
OPERATING GONDII

SYSTEM
"IONS

SRT
(days)

SECONDARY CLARIFIfRS
HYDRAULIC LOADING

OVERALL PAH
REMOVAL

OVERALL PCS
REMOVAL

OVERALL METAL
REMOVAL

HRT
(h) -

Bon

LOADING
(g/m3.h)

NEW SECTION AVG.

HYDRAULIC LOADING

(m3/m2-d)

OLD SECTION AVG.

HYDRAULIC LOADING

(m3/m2-d)

RANGE

AVERAGE

(%)

RANGE
AVERAGE

RANGE

AVERAGE

(%)

n'"C£MBER

JAMUARY

FEBRUARY

~w"fNTE^
AVERAGE

2.06
2.65

2.47

3.17
3.36

3.14

3.65
3.83
3.75

33.0
27.2
27.0

3.57

8.1

4.5

11.0

9.3

24.6

48.9
33.1
23.8

93 - 100

74 - 99

60 - 94

29.1

5.4

15.0

35.3

MARCH
APRIL
MAY

SPRING
AVERAGE

JUNE
JULY
AUGUST

26.1

14.2

17.6
^_

19.3

2.9
6.3

5.6

19.5
16.0
21.9

26.9
31.7
25.8

87 - 100

99 - 100

71 - 96

4.9

19.1

28.1

100

13.8

7.3

11.1

4.7
6.5
8.8

20.0
17.4
15.4

23.3
22.2
25.8

94 - 100

71 - 100

52 - 97

SUMMER
AVERAGE

3.74

10.7

6.7

17.6

23.8

OVERALL
AVERAGE

"overall

MAXIMUM

3.12
3.83
2.06

19.7
33.0

5.7
8.8

17.2

29.0

24.6

48.9

100

ICO

OVERALL
MINIMUM

7.3

2.9

9.3

22.2

- 405 -

old section clarifiers were hydrauHcally overloaded (23.8-48.9 m-^/m^d) com-
pared to the new section clarifiers {9.3-24.6 m-^/m^d). During the sunmer a
better flow split was achieved and the average hydraulic loadings were 17.6
m^/m^d In the new section and 23.8 m^/m^d in the old section.

In general, the aeration section was operating under both higher
organic loading (BOO 29.1 g/m^*h) and hydraulic loading during the winter
period than during the spring or summer periods. The most favourable average
conditions occurred during the summer period (highest HRT of 3.74 h; lowest
BOO loading of 10.7 g/m^-h; highest SRT of 6.7 days and lowest, most even
loading to the final clarifiers).

Despite the significantly higher loadings and the imbalance between
the old and new sections of the plant during the winter period the overall
treatment efficiency for the organics and metals studied did not differ sig-
nificantly between the seasonal periods. For example, the average removal
for the group of PAHs was 98^ In both the winter and summer periods. Simi-
larly for PCBs and metals there was no apparent correlation between the aera-
tion section operating conditions and removals of the contaminants, as very
■good removals were achieved throughout the study (e.g. 88-100^ PCB removals
and 31-90^ for the totalized group of seven metals).

From the results of this study it is apparent that municipal WPCP's
operating at 3-8 days SRTs, even at relatively low HRTs (2.5 to 4 h) and high
organic loadings (30 g BOO/m-^'h) should achieve excellent removals of trace
organics and metals. These results support the findings of various EPA and
Environment Canada studies, discussed by Melcer (1982), which generally ack-
nowledged that SRT is one of the most important factors affecting the biolo-
gical removal of priority pollutants.

Relative Removals by Prlaary and Secondary Treatment

Although the primary effluent was not analyzed for hazardous con-
taminants in this study, it has been found by EPA (1982) and others that a
correlation exists between suspended solids removal efficiency in the primary
clarifiers and trace organic and trace metal removals. Table 10 presents
estimated HC removal s for the primary and secondary sections of the WPCP,
which were calculated based on the assumption that the solids fraction of the
;,Cs would be removed in association with the suspended solids. In addition

lf6\.E 10. COt^WISON Of" MKll^rtr W€ SHCOJDARY REMOVALS OF tlAZARDOUS CONTAMINANTS

—

ToiM. IN

rLuLNl

SuLlO FKAOnON

LIvHl- ^HACTION

CALCULATEU

PRIH^RY EFFLUENT

MEASURED

FINAL

CALCULATED

MEASURED

EFFLUFNT 1

PERCENT

OVERALL

CX)H-i ' N

MASS

CONG , N

MASS

COfJC ' N

MASS

MINANT REMOVAL

CONC'N'

MASS*

CONC'N

MASS

REMOVAL IN

PERCENT

luy/DT

(k j/j)

(ug/L)

(kg/ J)

(ug/L)

(kg/d)

IN PRIMARIES*

(ug/L)

(kg/d)

(ug/Dt

(Kg/d)

SECONDARIES

REMOVAL

Noptitlialene

l^.i

4.4

11.0

3.2

4.2

1.3

8.5

2.5

0.3

0.08

Acenaphthy leiio

6. J

i.e

2.5

0.7

3.8

1.1

4.6

1.4

0.05

0.02

100

Dlbenzofuran

12.2

3.5

6.4

1.9

5.8

1.7

8.3

2.4

0.14

0.04

f 1 uor wne

16.1

4.7

9.4

2.7

6.7

1.9

10.4

J.t

0.22

0.06

F luorantheno

44.0

12.6

ib.9

10.3

8.1

2.3

22.1

6.3

0.7

0.2

Carbdzole

22.5

6.5

8.6

2.5

13.7

4.0

17.1

5,0

0.4

0.11

Pyrane

39.9

11.5

33.4

9.6

6.4

1.9

19.4

5.6

0.9

0.26

Benzo(a)pyrene

44.7

13.2

35.0

10.3

9.7

2.9

23.4

6.9

0.7

0.21

PAH Average

Removals

Total PCBs

0.131

0.04

0.084

0.03

0.047

0.01

0.077

0.022

0.025

0.01

Pentochlorophenol

0.25

0.07

0.02

0.004

0.21

0.006

0.216

0.06

0.1

0.05

Iron

6251

1817

4985

1464

1266

353

3210

924

478

135

Aluminum

19^

595

U!0

401

655

194

1166

350

376

Chromium

193

133

110

5.5

Copper

129

105

6.0

Nickel

Lead

Zinc

431

131

281

150

260

Metal Average

■*

• Calculated using average 61| SS removal In the prUarles and assuming SS removal - removal of contaminant soIU* fraction,
t These values differ some-hat from overall study averages reported in Table 4 because a fe« anomalous values which differed
from averages by an order of magnitude -ere omitted for purposes of this assessment.

- 407 -

It was necessary to assume that none of the dissolved fraction of the conta-
minants would be removed during primary treatment. Therefore, using the
average primary suspended solids removal on sampling days during the study
(i.e. 61%) in combination with contaminant phase loading data. It was possi-
ble to calculate the resultant primary effluent HC masses and concentrations.

Comparison of the calculated primary effluent values to measured
Influent contaminant loadings allowed the estimation of percent removal effi-
ciencies for the HCs In the primary clarifiers.

As can be seen in Table 10, the overall removal efficiencies calcu-
lated (e.g. 98% PAHs; 83% for the group of seven metals found in the greatest
quantities) are very close to the measured values (97% for the PAHs; 85% for
the metals), thus suggesting that the initial assumptions were valid.

Percent removals calculated for the PAHs in the primary clarifiers
ranged from 24% for acenaphthylene and carbazole to approximately 50% for
fluoranthene, pyrene and ben2o(a)pyrene. The overall average removal of PAHs
in the primaries was 3ft%. These results are slightly lower than results of
the U.S. EPA MERL pilot plant study (EPA, 1982), 1n which it was found that
56% of the influent PAH loading was removed through primary treatment. The
results in Table 10 are also lower than Petrasek's (1983) findings that up to
65% of the influent PAH mass may concentrate In the primary sludge. However,
the PAH removal efficiencies estimated support the theory that the trace
orgam'cs tend to adsorb onto the solid fraction of the wastewater matrix, and
thus, to a large extent are settled out with the SS in the primary clari-
fiers.

Average removals of PCBs in the primaries was estimated to be 41%
which was similar to the average of 50% Shannon (1976) found in his study of
33 municipal wastewater treatment plants in Ontario. Estimated PCP removal
in the primary clarifiers was only 5% due to the very small mass associated
with solid fraction of the wastewater.

The range for metal removal efficiencies in the primary clarifiers
was similar to the trace organics (e.g. 31% for nickel to 57^ for lead) due
mainly to the wide range of solubilities and other characteristics of the
metallic compounds studied. The average metal removal efficiency in the pri-
iTiaries calculated (44%) agreed well with the findings of Nomura (1974) who
reported that the percent removals of metals in the primary section of a
municipal wastewater treatment plant ranged from 14% for ni^.kel to 50% for

- 408 -

copper, averaging 41% overall. Oliver and Cosgrove in their 1975 stu(ly found
primary clarifier removals ranging from 15% for nickel to 69% for aluminum
with an overall average of 57%.

In summary, the data In Table 10 suggests the following trrnd? *-j-
gardlng primary versus secondary treatment efficiency for the HCs monitored.

i) PAHs tended to be removed to a greater degree In the secondary sec-
tion (average = 59% compared to average = 33% in primaries).

11) PCB removals in the primaries were estimated to be approximately
equal to those in the aeration section.
11i) Trace metals removals appeared to be slightly higher in the primar-
ies than in the secondary section.

iv) Relative removals of specific contaminants In the primary and
secondary sections of the plant appeared to vary mainly according
to the influent concentration distribution between solids arH
liquid phases (i.e. individual contaminant characteristics) . Both
primary and secondary treatment processes were essential to achieve
the efficient removals of HCs observed in this study at the Hamil-
ton WPCP (85-95% for most contaminants).

Effect of In-Pi ant Return Streaw Loadings

Duri ng thi s study , the i n-pl ant return stream was compri sed of
waste activated sludge, digester supernatant, vacuum filter filtrate, incin-
erator ash quench water, miscellaneous cleanup waters and periodic discharges
resulting from digester clean out or aeration basin emptying. Sampling was
limited to the combined influent, effluent, WAS (all sampling days) and the
combined in-plant return stream, including WAS (on selected days only).

In order to assess the effect of in-plant return stream loadings on
contaminant removal In the Hamilton WPCP, an extensive analysis was done to
determine the relative contributions of raw municipal sewage, combined in-
plant returns and the WAS return stream to the total combined influent.

A summary of the mass distributions for eleven of the trace organ-
ics monitored during this study is presented in Table 11. As discussed pre-
viously, the other pesticides measured were below detection limits in most
samples and were not included in this assessment. In addition to show'ng the

TABLE II. WSS DISTRIBUTIW OK TRACE ORGANICS BETWELN COfBINED INFLUENT. HAMILTON RAW SEWAGE,

TOTAL IN-PLANT RETURN AND WAS RETURN

COtJTAM 1 NANT

COMBINED

INFLUENT •

HAMILTON RAW SEWAGE

TOTAL IN- PL AN!

RETURN

WAS RETURN

SOLID
FRAC.

LIO.
FRAC.

TOTAL

t IN

SOLID

SOLID
FRAC.

LIO.
FRAC.

TOTAL

% IN

SOLID

i TOTAL
INFLUENT

SOLID
FRAC.

LIQ.
FRAC.

TOTAL

t IN

SOLID

t TOTAL
1 NFLUENT

SOLID
FRAC

Lig.

FRAC.

TOTAL

i IN

SOLID

% TOTAL
RETURNS

t TOTAL
1 ffLUEHT

Naphthalene

0.4

1.201

1.60

0.3

1.2

1.5

0.1

0.001

0.101

0.005

100

0.3

Acenaphthy lene

0.5

0.705

t.2l

0.4

0.7

1.1

0.1

0.005

0.105

0.005

100

0.4

Dibenzofuran

0-5

0.513

I.OI

0.4

0.5

0.9

0.1

0.013

0.113

0.008

0.006

100

0.8

F 1 uorene

0.7

0.52

1.22

0.5

1.0

0.2

0.02

0.22

0.006

100

0.5

F luoranthene

2.7

0.429

3.13

1.5

0.4

1.9

1.2

0.029

1.229

0.157

100

5.0

Carbazole

1.9

2.822

4.72

1.4

2.8

4.2

0.5

0.022

0.522

0.265

100

5.6

Pyrene

2.5

0.423

2.92

1.0

0.4

1.4

1.5

0.023

1.523

0.213

100

7.3

Benzo(alpyrene

7.6

3.251

10.85

5.8

3.1

8.9

1.8

0.151

1.951

0.246

100

2.3'

Average of PAHs

2.a

Lindane

0.0

PCS

0.03

0.01

0.04

Pontachlorophenol

0.004

0.06

0.064

T
-- — ~

Notes : All loadings In Kg/d

T » Trace
NA =■ Not Available; too low to allow calculation.

• tntluent values In this table differ from those In Table 10 because data Is from
14 days on Table 10 and 8 days this table (returns only sampled on 8 days).

- 410 -

relative percent of the total influent contaminant masses for each compound,
Table 11 shows the breakdown of these masses between the liquid and solid
phase of each source.

Because of the extremely low concentrations of lindane, PtBs ar.j
PCP, particulary in the return streams, accurate estimates for thise com-
pounds could not be developed. It is noted however, that tracer of PCB and
PCP were detected in the in-plant return and WAS return, whereas, no lindane
was found in the return streams.

With regard to the PAHs, a combined average of 80^ of the mass was
from the Hamilton Sewage (50% associated with the solids), meaning that as an
overall average. 20% of the PAHs monitored came from the in-plant return
stream. The contribution of PAHs in the in-plant return stream ranged from a
low of 6X for naphthalene to a high of 52% for pyrene.

It is Important to note that the percentage of PAHs in the in-plant
return solids fraction was consistently high for all compounds (i.e. 88% -
99%, averaging 95%). On the other hand, there was considerable variability
in the amount of each PAH associated with the solids fraction of the Hamilton
sewage. The general trend observed was that for contaminants which tended to
be concentrated in the solids fraction in all samples, a relatively high per-
centage of the total contaminant loading came from the in-plant return stream
(e.g. fluorene - 18%; fluoranthene - 39%; pyrene - 52%; and benzo{a)pyrene -
18%), and for these compounds the contribution from the WAS return was also
higher {e.g. fluoranthene 5% of total Influent mass; carbazole 5.6%; pyrene
7.3"). The solid fraction of the WAS return stream accounted for essentially
100'^ of the WAS stream organics, because extremely low concentrations were
measured in the liquid phase.

Since the contaminant contribution of the WAS stream to the in-
plant return stream was relatively small for most of the organics monitored
(5-14% for all but carbazole at 51%). It 1s apparent that other streams mak-
ing up the total in-plant return were making significant contributions to the
total influent loading. This observation, and a review of the results of a
study by CANVIRO (1984) in which the fate of PAHs and metals during anaerobic
digestion was Investigated, point to the digester supernatant as a major
source of some PAHs. For example, the two PAHs with the greatest contribu-
tion ""rom the in-plant returns in this study were fluoranthene and pyrene.
In the CANVIRO (1984) study, fluoranthene and pyrene solubllized to the

- 411 -

greatest extent during digestion (approximately 300% for both), and these two
compounds were present In the raw Hamilton digester sludge in concentrations
higher than any of the other PAHs monitored In that study (6-10 mg/L aver-
age).

Table 12 provides comparisons of the mass distributions within the
plant for the trace metals group. The elements arsenic, cadmium, mercury and
selenium were not present 1n the influent or return streams In concentrations
large enough to provide any meaningful results or to establish any trends.
Thus, discussion of the metals distribution Is limited to 7 of the metals
(Fe, Al. Cr, Cu, Ni . Pb, Zn).

Trace metal loadings in the solid fraction of the combined influent
ranged form 2040 kg/d for iron to 13.5 kg/d for nickel. Average liquid frac-
tion loadings for the combined Influent varied from >440 kg/d for Iron to
<3.0 kg/d for lead. Trends in metal loadings for the combined Influent indi-
cate that as an overall average approximately 73% of the total mass was con-
tained in the solid fraction of the wastewater. This Is consistent with
Kang's (1981) assertion that the majority of metallic contaminants present in
a sewage treatment plant will exist in a solid form because of the relative
insolubility of the metals.

The results in Table 12 show the Hamilton sewage Input as contribu-
ting approximately 79o of the overall metals loading to the plant. Thus, an
overall average of 2U of the total metals mass entering the plant resulted
from the in-plant returns. Unlike the organics there was little difference
in the percent contribution of the in-plant return from metal to metal, with
the exception of the more soluble nickel (12% from the in-plant return).
Within the total return stream the metals were consistently associated mainly
with the solids fraction (i.e. 89-97% for the 7 metals; average 95%). Mass
loadings in the solids fraction of the total in-plant return ranged from 3.2
kg for nickel to 491 kg for iron, while In the liquid phase, values were from
0.2 kg for lead to 28 kg for iron.

The WAS contribution of heavy metals to the total return flow was
substantially greater than for organics. as an average 39% of the return
metals mass was contained in the waste activated sludge. This trend was very
consistent from metal to metal during the study as Indicated by the range of
35%-43% of various metals coming from the in-plant return. The WAS return

TAbLt 1^. MASS UlStKlbUIlON OF fHACt NtfALS BETWEEN COMUINED INFLUENT, HAMILTON RAW SEWAGE

TOTAL (f+-PLANT RETURN AND WAS RETURN

CON T AMI MAN T

COMBINED

INFLUENT •

HAMILTON RAW

SEWAGE '

TOTAL m-PLAN1

RETURN

WAS RETURN

SOLID
FRAC.

LIQUID
FRAC.

TOTAL

i IN

SOLID

SOLID
FRAC.

LIQUID
FRAC.

TOTAL

t IN

SOLID

% TOTAL
INFLUENT

SOLID
FRAC.

LIQUID
FRAC.

TOTAL

i IN

SOLID

1 TOTAL
1 NFLUENT

SOLID
FRAC.

LIQUID
FRAC.

TOTAL

t IN

SOLID

% TOTAL
RETURNS

i TOTAL
INFLUENT

Iron

2040

442

24B2

1549

414

1963

491

519

204

0.6

204.8

Aluminum

431

198.5

629

299

194

493

131

4.5

135.5

46.6

0.6

49.2

Chromium

44.6

19.7

30.4

19.1

49.5

14.2

0.6

14.6

6.1

6.14

Copper

33.1

8,9

24.1

8.6

32.7

9.0

0.3

9.3

3.9

NIcket

13.5

15.7

29.2

10.3

15.3

25.6

3.2

0.4

3.6

1.5

Lead

23.8

2.9

26.7

17.2

2.7

19.9

6.6

0.2

6.8

2.4

Zinc

39.5

110.5

38.5

87.5

1.0

11.5

0.1

11.6

Average of 7
Meta 1 s

380.0

103.4

483.4

282.7

123.4

381.5

96.1

5.0

t03.1

39.7

0.2

39.9

40.4

Notes: AM loadings In kg/d

» Influent values In this table differ from those In Table 10 because data Is from
14 days on Table 10 and 8 days this table (returns only sampled on 8 days).

t— 1
I

- 413 -

stream also made a significant contribution to the metals loading In the com-
bined influent, as It made up 8% of the total metal mass. Similar to the
trace organlcs, the trace metals concentrated almost exclusively in the solid
fraction of the WAS with less than 0.5% of the metal load being attributable
to the liquid fraction,

A review of the results from Individual sampling days, although not
conclusive, showed a trend toward higher percentages of metals being returned
in the winter and spring months. For example, loadings for Fe and Al In the
In-plant returns peaked In the winter months, while Cr, Cu, N1 , Pb and Zn
concentrations were greatest In the spring. Further study Is required to
verify this trend and further identify the sources of contaminants.

CONCLUSIONS AND RECOmENDATIONS

Conclusions

The following conclusions relating to the removal of hazardous con-
taminants in the Hamilton WPCP can be drawn from this study.

1) Concentrations of contaminants entering the Hamilton WPCP were
within the ranges found in previous U.S. EPA and Environment Canada
studies of municipal treatment plants serving Industrialized areas.

2) Results indicate a high degree of overall removal (>97% for the
WPCP as a whole) for the PAHs. Total PCBs were similarly well re-
moved, averaging 90%. Both lindane and pentachlorophenol were re-
moved to a lesser extent (70% and 63%, respectively) and with con-
siderably less consistency. Trace metals removals were generally
in excess of 80% (overall average = 85%, ranging from 62% for
nickel to 95% for lead).

3 ) Annual 1 oadi ngs of HCs enteri ng and bei ng di scharged from the
Hamilton WPCP were estimated as follows:

- 414 -

TRACE

ORGAN I CS *

TRACE METALS '

CONTAMINANT

OUT

CONTAMINANT

OUT

(kg/yr)

Naphthalene

1431

30.3

Iron

742,410

51,027

Acenaphthylene

613

4.1

Aluminum

250,755

39,311

Dibenzofuran

1150

12.4

Arsenic

219

Fluorene

1526

19.0

Cadmium

110

Fluoranthene

4066

65.7

Chromium

83,111

2,008

Carbazole

2310

43.8

Copper

14,491

2,026

Pyrene

3705

84.0

Mercury

29.2

Benzo(a)pyrene

4420

69.4

Nickel

9,892

3,979

Lindane

7.3

3.7

Lead

10,585

694

Total PCBs

14.7

4.0

Selenium

183

Pentachl orophenol

23.4

11.0

Zinc

346.385

10.038

* These estimates are based on average values from 14 sampling days and
Include ^^ery high values measured on one or two days for most contami-
nants. Thus, these annual averages are considered to be high estimates.

4) Despite experiencing significantly higher loadings and poorly bal-
anced loadings between the old and new sections of the plant during
the winter and spring periods, the overall treatment efficiency for
the organics and metals studied did not differ significantly be-
tween the seasonal periods (e.g. winter and summer average PAH re-
movals were both 98%; 88 to 100% for PCBs; 81 to 90% for totalized
metals).

5) From the results of this study, It is apparent that municipal
WPCP's operating at 3-8 days SRTs, even at relatively low HRTs (2.5
to 4 h) and high organic loadings (30 g B0D/m3.h) can achieve
excellent removals of trace organics and metals. This supports the
belief that SRT is one of the most important factors affecting
biological removal of priority pollutants.

6) The following trends were Indicated regarding primary versus secon-
dary treatment efficiency for the HCs monitored:

i) PAHs tended to be removed to a greater degree in the secondary
section (average = 59* compared to average = 38% in primar-
ies),
ii) PCB removals in the primaries were estimated to be approxi-
mately equal to those in the aeration section,
iii) Trace metals removals In the primaries were slightly higher
than in the aeration section.

- 415 -

1v) Relative removals of specific contaminants in the primary and
secondary sections of the plant appeared to vary mainly accor-
ding to the influent concentration distribution between solid
and liquid phases (i.e. individual contaminant characteris-
tics). Both primaT7 and secondary treatment processes were
essential to achieve the efficient removals of HCs observed in
this stu4y at the Hamilton WPCP.

7) Assessment of the contaminant loadings originating from the raw
Hamilton sewage, the total in-plant return stream (including WAS)
and the WAS return stream showed that as an overall average of all
the HCs monitored, approximately 20% of the loading originated from
the In-plant return stream. This varied considerably between spe-
cific contaminants, being less than 10% for some PAHs, 52% for
pyrene, 12% for nickel and 20-25% for the other metals.

8) The WAS return stream contributed overall averages of 14% and 39%
to the total in-plant return PAHs and metals loadings, respective-
ly. Thus, the WAS return represented on average 3% and 8% of the
total influent loadings of PAHs and metals, respectively. Only
trace amounts of PCBs and pesticides were detectea In the return
stream.

9) It was found that both the organics and metals in the total in-
plant return stream were consistently associated almost exclusively
with the solids fraction (>95% average). This was in contrast to
the wide range of percentages found in the raw sewage (e.g. 20%-79%
for PAHs and 40%-86% for metals).

10) Since only 14% of the PAHs and 39% of the metals in the total in-
plant return were accounted for by the WAS return stream. It is
apparent that other In-plant returns such as digester supernatant
and vacuum filter return must be contributing significant quanti-
ties of contaminants to the combined WPCP influent loadings.

- 416 -

Recoiendatlons

Recoanendatlons relating to the operation of the Haallton WPCP In-
clude:

1. Continued efforts should be made to maintain an even balance of
loading between the old and new sections of the plant

2. The plant should be operated at a minimum 5 day SRT to maximize re-
moval of HCs entering the plant.

3. SRT control should be practiced separately on the old and new sec-
tions.

4. Action should be considered to minimize the suspended solids con-
tent of the in-plant return streams (other than WAS) since over 95%
of the contaminants In the total In-plant return were associated
with the solids fraction.

5. Industrial sources of the shock loads of certain contaminants, par-
ticularly zinc should be identified and eliminated since occasional
daily average concentrations observed in this study were in the
range known to be inhibitory to the activated sludge process.

Reconnendatl ons for further research arl si ng from thi s study 1 n-
clude:

1. Investigations should be carried out to verify the percent contri-
butions of the various in-plant return streams » as approximately
80% of the PAHs and 60% of the metals in the total return stream
are not accounted for by the WAS return.

2. Since overall seasonal and other trends were difficult to establish
because of the limited number of samples collected in this study,
the methodol ogy shoul d be uti 1 i zed under more control 1 ed condi -
tions, such as sampling on days when certain known events are
occurring rather than on randomly selected days.

3. Emphasis should be placed in future studies on the factors affec-
ting primary treatment efficiency and on sludge handling process
factors contributing to the trace organics and metals removals/re-
cycle within municipal WPCP's.

- 417 -

REFERENCES

1. Bishop, D.F. et a1 , "Control of Specific Organic and Metal Contami-
nants by Municipal Wastewater Treatment Processes." Municipal
Environmental Research Laboratory, U.S. Environmental Protection
Agency, Cincinnati, Ohio, 1982.

2. Bridle, T.R. and B.E. Jank, "Removal of Trace Organlcs by Biologi-
cal Treatment". Short Course on the significance. Analysis and
Control of Toxic Organic Substances in Wastewater, Edmonton, Al-
berta, Canada, 1980.

3. Bridle, T.R., "The Impact of Hazardous Organlcs on Sludge Manage-
ment and Disposal". Presented at the PCAO/MOE Seminar Hazardous
Substances In Wastewaters, November 3, 1982, Toronto, Ontario.

4. Brown, H.G. et al , "Efficiency of Heavy Metals Removal in Municipal
Sewage Treatment Plants". U.S. Environmental Profectlon Agency
Environmental Letters, 5(2), 103-114, Kansas City, Missouri. 1973.

5. CANVIRO. "Detailed Review of Thirty Municipal Wastewater Treatment
Facilities in the Great Lakes Basin". Prepared for Work Group III,
IJC, 1983.

6. CANVIRO, "Final Report on Sludge Processing Operations on the Fate
and Leachability of Toxic Contaminants 1n Municipal Sludges". Sub-
mitted to Wastewater Technology Centre, EPS and Environment Canada,
1984.

7. Chaney, R.L., "Health Risks Associated with Toxic Metals in Munici-
pal Sludge", in " Sludge - Health Risks of Land Application ", Edited
by G. Bitton, B.L. Damron, G.T. Edds, and J.M. Davidson. ^Ann Arbor
Science, pp 59-83, 1980.

8. CMA/EPA, 1982, "CMA/EPA Five-Plant Study". Prepared for Chemical
Manufacturers Association by Engineering-Science Inc., 3109 North
Interregional, Austin, Texas 78722.

9. Cohen. J.M. et al , 1981, "National Survey of Municipal Wastewaters
for Toxic Chemicals". MERL. U.S. EPA, Cincinnati, Ohio 45268.

10. Craig, G. et al , 1980, "Survey of Nine Ontario WPCP's for Organic
Trace Contaminants", Water Resources Branch, MOE, Report In prepar

11. Dacre. J.C. 1980. "Potential Health Hazards of Toxic Organic Resi-
dues in Sludge", In " Sludge - Health Risks of Land Application ".
Edited by G. Bitton, B.L. Damron, G.T. Edds, and J.:^. Davidson.
Ann Arbor Science, pp 85-102.

- 418 -

12. Daniel, F.B. et al . 1979, "Biochemical Studies on the Metabolism
and DNA-Binding of DMBA and Some of its Monofluoro Derivatives of
Varying Carcinogenicity", in "Polynuclear Aromatic Hydrocarbons".
Edited by P.W. Jones and P. Leber. Pub. Ann Arbor Science.

13* Environment Canada, "Sources of Metals and Metal Levels In Munici-
pal Wastewaters". Research Report No. 80, 1978.

14. EPA, "Fate of Priority Pollutants in Publicly Owned Irtuu.ient
Works". Final Report. EPA 440/1-82/303. September 1982.

15. EPA, "Fate of Priority Pollutants in Publiciy Owned Treatment
Works". 30 Day Study. EPA 440/1-82/302, July 1982.

ation.

16. Holzclaw, P.W. and M.D. Neptune, "Approach of Qaulity Assurance/
Quality Control In the Organic Chemicals Industry Monitoring Pro-
gram", J. Environ. Sci . Health. A15:5. pp 525-543, 1980.

17. Jenkins. D.I. and L.L. Russell, "Impact of Priority Pollutants on
publicly Owned Treated Works Processes. A Literature Review."
Source unknown.

18. Jones, P.W. and P. Leber, 1979, " Polynuclear Aromatic Hydrocar-
bons". Third International Symposium on Chemistry and Biology -
■Ctrxinogenesis and Mutagenesis. Pub. Ann Arbor Science, 1979.

19. Kang, S.J., J.W. Bulkkey and J.L. Spangler, "Fate of Heavy Metals
and Tolerance Limits In POTW", in Proc. of the ASCE 1981 National
Conferen ce on Environmental Engineering. Edited by P.M. Saunders,
Pub. ASCE. 1981.

20. Melcer, H. 1982, "Biological Removal of Organic Priority Pollu-
tants". Presented at the Hazardous Substances in Wastewaters Semi-
nar sponsored by the Pollution Control Association of Ontario and
the Ontario Ministry of the Environment. November 3, 1982, Tor-
onto.

21. Ministry of the Environment (MOE) Ontario, "Water Management" {also
known as MOE "Blue Book"), Toronto, 1978.

22 . MOE . "Gui del i nes for the des 1 gn of Sewage Treatment Works . "
Ontario, 1980.

23. Munro, J.R. et al . 1982, "A Survey and Evaluation of Organic Com-
pounds in Nine Sewage Treatmemt Plant Effluents in Southern
Ontario". Prepared for EPS and HOE, internal report.

24. Nelson, P.O., A.K. Chung and M.C. Hudson, 1981, "Factors Affecting
the Fate of Heavy Metals in the Activated Sludge Process". Journal
WPCF. Vol. 53, No. 8, pp 1323-1333.

- 419 -

25. Nomura, M.M, and R.H.F. Young, "Fate of Heavy Metals In the Sewage
Treatment Process". Water Resources Research Center, University of
Hawaii, Technical Report No. 82, 1974.

26. Oliver, B.G. and E.G. Cosgrove, "The Efficiency of Heavy Metal Re-
moval by a Conventional Activated Sludge Treatment Plant". Depart-
ment of the Environment, Burlington, Ontario, Cananda, 1973.

27. Patterson, J.W. and P. Shimada and C.N. Haas, "Heavy Metals Trans-
port Through Municipal Sewage Treatment Plant". Department of the
Environment, Burlington, Ontario, Canada, 1973.

28. Petrasek. A.C. et al , "Behaviour of Selected Organic Priority Pol-
lutants in Wastewater Collection and Treatment Systems", presented
at 53rd Annual WPCF Conference, Las Vegas, Nevada, September 1980.

29. Rush, R.J. and L.J. Taylor, "Removal of Hazardous Contaminants
(HCs) in an Ontario Water Pollution control Plant (WPCP)". Presen-
ted at 5th Annual Technology Transfer Conference, Toronto, Canada,
Nov. 1983.

30. Shannon, E.E., H.D. Monteith and A.K.W. Ho, "Monitoring of Selected
Trace Organics Duri ng Biol ogical Wastewater Treatment Systems" ,
presented at 53rd Annual WPCP Conference, Las Vegas, Nevada, Sep-
tember 1980.

31. Strier, M.P. and J.D. Gallup, "Removal Pathways and Fate of Organic
Priority Pollutants in Treatment Systems: Chemical Considera-
tions". U.S. E.P.A.. Washington. O.C., 1982.

32. Thakker, D.R. et al , "Comparative Metabolism of a Series of Poly-
cyclic Aromatic Hydrocarbons by Rat Liver Microsomes and Purified
Cytochrome p-450", in "Polynuclear Aromatic Hydrocarbons". Edited
by P.W. Jones and P. Leber. Pub. Ann Arbor Science, 1979.

33. van Rensburg J.F.J, et al , 1980, "The Fate of Organic Micropollu-
tants Through an Integrated Wastewater Treatment/Water Reclamation
System". Prog. Water Tech.. Vol. 12, Toronto, pp 537-552, 1980.

34. Zedeck, M.S.. 1980, "Polycyclic Aromatic Hydrocarbons - A Review".
J. of Environmental Pathology and Toxicology. 3, pp 537-567, 1980.

- 421 -

ASSESSING THE IMPACT OF HAZARDOUS IMMISCIBLE LIQUIDS IN SOIL

G.J. Farquhar and E.A. McBean
University of Waterloo

November 1984

- 422 -

ABSTRACT

This paper describes the current status of research assessing the
behavior of imniiscible hazardous contaminants in soil. The work. Is
supported in part by Provincial Lottery Funds and is being performed Jr
the Department of Civil Engineering at the University of Waterloo.

A major portion of the research involves the simulation of
immiscible liquid behavior in soil through the use of computer models.
This includes (i) a model for spills onto soil with overland flow,
penetration and evaporation, (ii) a model for the transport of
immiscible liquids in soil either saturated or unsaturated, and (iii) a
model for the movement of hazardous vapours in unsaturated soil.

Another area of research is experimental in nature but is only in
Its initial stages. It involves (i) the development of methods to
detect immiscible liquids in soil through measurements of dielectric
coefficients, thermal conductivity and electrical conductivity, and (ii)
the performance of experiments to yield information on the flow of
immiscible liquids in soil, especially the relationships between
relative permeability, capillary pressure and percent saturation.

Research is also in progress to develop a spill response model
which incorporates the behavior of the spilled liquid including overland
flow, penetration, evaporation and transport within the soil and also
recommends remedial action specific for the incident.

A literature review on the subject is near completion. It has
approximately four-hundred references and was prepared with the
assistance of a key-word information storage and retrieval system
developed on a microprocessor for this research.

- 423 -

INTRODUCTION

In recent years, increased emphasis has been placed on hazardous
materials spilled onto soil and the impact of these materials on the
soil and groundwater environment. The transport of the contaminants,
their Interactions with the solid, aqueous and gaseous phases of the
subsurface and the development of remedial measures have been areas of
major research effort.

Immiscible liquids such as petroleum products and solvents are of
special concern because many are defined as haaardous and all have the
property of not mixing with water. This property adds greatly to the
complexity of these contaminants as they penetrate the soil. Although
the behavior of soil water Is reasonably well understood, the behavior
of multiphase fluids consisting perhaps of an aqueous phase, an
Imniscible liquid phase and a gas within a solid soil phase is not. It
is because of this need for additional information together with the
frequency and hazard of immiscible liquids spilled onto soil that this
work was undertaken.

The research has been divided into four major sections comprised
of:

1. Literature Review

2. Simulation Model Development

3. Experimentation

^. Spill Response System Development

The schedule for the research has been set for a three-year period,
of which the first year is nearing completion. Funding has been
provided by the Ontario Ministry of the Environment from Provincial

- 424 -

Lottery Funds and the Natural Sciences and Engineering Research Council.

This document Is a progress report describing the work which has
been accomplished to date-
LITERATURE REVIEW

To this point in the research, approximately 400 documents and
papers related to hazardous immiscible liquids and their behavior when
spilled onto soil have been reviewed. Two computer-based search systems
WATMARS and AWWA were used to identify relevant information.

An information storage and retrieval package (ISRP) using the IBM
DBASE II Software System was developed as a part of this work. The ISRP
has the following characteristics,

1 . Storage Format

This includes reference number, authors, title and publication
data stored on 5-1/4" floppy disk.

2. Key Words

Each document is assigned up to 10 sets of key words from the
list shown in Table 1.

3. Comment

A comment about the document and its content is stored on the
disk.

4. Retrieval

The Users' Manual for the ISRP describes how the system is
accessed through use of key words. The user is able to print
out the author, title and comment for all documents stored
with the key words entered.
The data stored can be added to or modified as more information is
acquired.

- 425 -

A report describing the results of the literature review is In
preparation. A first draft has been completed. The Table of Contents
of the report is shown in Table 2.
SIMULATION MODEL DEVELOPMENT

Very little quantitative information is available on the behavior
of hazardous immiscible liquids in soil. This applies to both field and
laboratory investigations. It is especially the case for
soil/contaminant interactions, relationships for permeabilities and
capillary pressures for variably saturated, multifluid systems and the
spreading and infiltration of spilled liquids onto soil. It also does
not appear that this kind of quantitative data will be available In the
near future. Consequently, it was decided that the development of
mathematical simulation models should be a part of this research in
order to study immiscible fluid behavior more extensively.

Mathematical models are simplified representations of physical
phenomena. However, within the constraints of the underlying
assumptions, these models can be used to examine a range of conditions
which would not be feasible to investigate on an experimental basis. In
this research it is also intended that a series of simulation models be
linked together to encompass a complete spill including spilling,
evaporation, penetration, transport in the unsaturated zone and
transport on and in the zone of saturation. The models will ultimately
be used both to learn about various spill conditions and to assess
various remedial strategies.

At this point in the research, work has been done on three
simulation models as discussed briefly below.

- 426 -
1. Spill Simulation Model (Wall, 1984)

The spill simulation model developed in this research takes account
of two specific conditions;

a. discharge from a source onto an inclined soil surface producing a
hydraulic bore of spilled material moving down the slope (analogous
to border irrigation), and

b. discharge onto a horizontal soil surface creating an enlarging
circular pool of spilled liquid.

In both cases, provision is made for variable liquid inflow,
Infiltration to the soil and evaporation Into the atmosphere. The
processes which govern spill formation have been expressed in the form
of the St. Venant Equations with rectangular co-ordinates for condition
"a" and polar co-ordinates for condition "b". These equations are
nonlinear, nonhomogeneous , first order, hyperbolic partial differential
equations. The system also accounts for supercritical and subcrltical
flow, a moving hydraulic jump and a moving bore. Because of the
complexity of these equations, analytical solutions were not possible.
Several approximate solution methods were examined in detail with the
MacCormick explicit finite difference scheme proving to be satisfactory
in most cases. The preservation of both mass and momentum and numerical
dispersion were checked as solutions were generated by the model.

The spill of a viscous volatile liquid Involves the dynamic
interaction between forces which drive the liquid over the soil surface
(momentum and pressure) and those which resist it (viscous and surface
friction).

Resistance to overland flow controls the depth, the shape of the
spill and the location of the hydraulic jump. Because of the presence
of fluid turbulence, resistance to flow in this model was expressed in

- 427 -

terms of the Darcy-Weisbach relations with Che introduction of a
constant friction factor (Kincald, 1970; Mlllel and Hornberger, 1979).

Infiltration of the liquid during spill formation is a complex
system which depends on both capillary and gravity forces. The approach
taken in this work is a simplification of the true system. A tubular
configuration with both capillary and gravity components was used-
Coefficients for the model must be determined experimentally.

Evaporation from the pool formed by the spill was represented as a
mass transfer equation (Sutton, 1953; MacKay and Matsugu, 1973) adapted
for southern Ontario conditions.

Other models have been developed to simulate a liquid spill but
these were judged to be insufficient for this work. In most cases the
authors failed to account for the energy in the inflow. Many» in fact,
used an instantaneous spill with constant depth. Others omitted
infiltration and evaporation from the analysis. The need for a more
complete model structure was recognized.

The literature review associated with this study failed to locate
existing data, from either field or laboratory experiments, for use in
verifying this spill simulation model. Attempts will be made in the
future to obtain or develop data on spill formation.
1. 1 Test Simulations

A series of test spill simulations were performed with the model;

a. to evaluate its overall performance, especially with respect to
mass and momentum transport,

b. to test the sensitivity of spill formation to model input
parameters such as rate of inflow, soil roughness and infiltration
rate, and,

- H<:o -

c- to examine the spilling of various liquids under a range of

conditions .

Two examples of the simulations performed are shown below.
Case 1

Case 1 simulated a slow spill of liquid from a long container,
(perhaps a train tank car on its side) onto a sloping soil surface*

- flow = 2.32x10"^ m^/s per m of tank width

- soil slopeCs) = 7.5xlO~^ m/ra

- soil roughness (n) = 3x10 " m

- node spacing = 2x10 m

- evaporation and infiltration were not considered.

The model used in this analysis was in the rectangular co-ordinates
format because of the sloped surface. The output is shown in Figures 1,
2, and 3. Figure 1 shows the location on the liquid front (or moving
bore) with respect to time. It must be noted that edge effects have
been excluded from the model . Therefore , no lateral spreading of the
liquid is accounted for, thus producing a rectangular-shaped spill. The
simulation becomes a "worst case" in terms of spill length being longer
but narrower than the real spill.

Figure 2 shows the fluid profile (depth vs. distance) at 5 time
period up to 600 seconds and compares uniform and critical depths for
the flow field (note vertical distortion). Figure 3 shows the velocity
profile. It can be seen that » as expected, the spill depth increased
with a corresponding reduced velocity as the spill distance increased.
Both fluid mass and momentum were preserved , Since subcritical flow
existed, no hydraulic jump was produced.

- 429 -

Case 2

Case 2 involved the spill of a liquid onto 3 flat surfaces
consisting of asphalt (Case 2a), clay loam (Case 2b) and coarse sand
(Case 2c). The source of the spill was assumed to be a ruptured
container with volume of 62,5 m , a rupture effective diameter of 0.4m
and a discharge coefficient of 0,85, The decreasing flow produced from
this configuration was Q " 0.704 - 0.004t with t in seconds. A specific
rate of infiltration was assigned to each soil (0 for asphalt and >0 for
clay and sand). The ponded liquid evaporation rates were the same for
each soil.

The model written in polar co-ordinates was used in this simulation
to produce a circular pond with time increasing to 326 seconds.

Figure 4 shows the depth (note vertical distortion) profile for the
spill on the asphalt surface. The relative locations of the spill front
and the hydraulic jump are shown.

Figure 5 shows the location of the spill front and the hydraulic
jump for the 3 surfaces with respect to time. The spill front on coarse
sand actually retreats somewhat in response to the combined effects of
decreasing flow, infiltration and evaporation.

Figure 6 shows the situation for the clay loam by comparing the
total spilled volume with the ponded. Infiltrated and evaporated
volumes. By comparing Figures 5 and 6, it can be seen that the pond
volume decreases although the pond diameter increases.
1 . 2 Summary

The results of the model testing to this point in the work indicate
Chat it can produce realistic spill simulations in response to a variety
of input conditions while conserving mass and momentum. Additional work,
will be done in the future ( 1 ) to streamline the model to reduce

- 430 -

computational ttrae, (2) to validate It If actual spill data can be found
or developed and (3) to Incorporate the model Into a spill response
model ■
2. Iimnlscible Liquid Porous Media Flow Model (Osborne, 1984)

The purpose of this work was to create a model to simulci:'_ the flow
of immiscible liquids through porous media after introduction to the
soil either from a spill or through direct disposal. While the model
has been essentially completed, linkage between it and the spill
simulation model has yet to be attempted.

Mathematical simulation of two-phase flow in porous media has been
used in the petroleum industry for some time to analyze the behavior of
gas and oil in reservoirs, and to optimize the effects of steam- and
water- flooding. Various numerical techniques. Including the finite-
difference and finite-element methods have been used to solve the
immiscible displacement equation. .However, application of such
numerical models Co immiscible groundwater contanvihation problems is a
relatively new practice.

In this work, a two-dimensional, two-phase mathematical model was
developed, based on Darcy's law and conservation of mass for each
liquid. The result was a pair of coupled, nonlinear partial
differential equations which display both parabolic and hyperbolic
characteristics, depending on the magnitude of a nonlinear coefficient.

A numerical model was developed to solve the equations using a
generalized method of weighted residuals in conjunction with the finite-
element method and linear quadratic Isoparametric elements. To
alleviate numerical difficulties associated with hyperbolic equations,
upstream weighting of the spatial terms in the model was Incorporated.

- 431 -

The theoretical and numerical accuracy of the model was verified by
comparison with simulation results with those from an existing one-
dimensional finite difference two-phase flow simulator (Little, Arthur
D, Inc. 1983). This form of verification was necessitated by the lack
of actual data on immiscible liquid transport in soil.

One of the comparisons made involved a vertical column of soil 7.0m

in length and consisting of the 3 soil components shown in Figure 7.

The initial distribution of non-aqueous phase liquid (NAPL) is also

shown plotted as elevation against X NAPL saturation (S ) within the

zone of saturation. Relationships between relative permeability (k ),
capillary pressure (P^) and % wetting fluid saturation (S ) were the
same for both models. However, the model developed here is not equipped
for hysteretic functions since its intended application is for the first
time displacement of water by the immiscible liquid. Thus, only the
primary drainage curve was used.

Figures 7 also shows comparisons between the NAPL profiles
predicted by the two models. The agreement was thought to be reasonably
good and was therefore taken as evidence that the model developed in
this work was performing in a satisfactory way.

The finite-element model was then used to simulate the migration of
an immiscible organic solvent in groundwater, from a chemical waste
dispersal site located north of Niagara Falls, New York. The effects of
uncertainty regarding several of the liquid and porous media properties
were examined, and it was concluded that the value of the model was
limited less by the numerical approximations involved than by the
accuracy of input parameter estimations. The results of this simulation
are presented in the work of Osborne (Osborne, 1984).

- 432 -

3. Vapour Transport Model

At locations where volatile liquids have been spilled or have
migrated away from disposal sites, concern often arises over the hazards
resulting from the vapours produced at the liquid interface. Both
experimental and theoretical work are being used to study this problem.
The experimental work is discussed subsequently. The theoretical work
will involve the development of a vapour transport model for soil
environments .

The vapour transport model has not yet been developed but will be
based on an existing model created by the authors to simulate the
movement of methane (CH^ ) gas through soil from landfill sites

■4

(Metcalfe, 1982; Farquhar and Metcalfe, 1982). The model is based on
equations for continuity and flow in two dimensions.

The flow or connective equation is the Darcy Equation. Estimates
of gas conductivity in variably saturated soil were used (Bear, 1972).
The continuity equations accounted for dispersion, convection and loss
terms. Gas diffusion in porous media, dispersion, dissolving from gas
into liquid phase and density differences between gases were taken Into
account. Discharge of gases at the soil surface was simulated using
transport equations for a laminar sub-layer (Thlbodeaux, 1979).

The CH^ transport model was tested using data collected at two
locations, the Mlssissauga Landfill and the Ottawa Street Landfill in
Kitchener, Ontario, The results from the Mlssissauga test are briefly
discussed below.

A plan view of the Mlssissauga Site is presented in Figure 8. The
location of a pumped (vacuum) gas venting system along the one side of
the landfill is shown as are the locations of piezometers in a line
perpendicular to the edge of the landfill. During the Fall of 1980, the

- 433 -
venting system was shut off for 85 days to permit landfill gas to
migrate outward into the adjacent soils. Gas pressures and
concentrations were measured at the piezometer locations during this
period.

A summarized soil profile along the plezomenter line is shown in
Figure 9. The progression of CH with time away from the landfill is
shown in Figure 9, 10 and 11 as CH percentiles (by volume) plotted vs.
time. The model simulations are also shown and indicate reasonably good
agreement between the actual and simulated concentrations . These
encouraging results support the extension of the model to simulate
vapour movement in soil.
EXPERIMENTATION

A series of experiments has been proposed for this study (a) to
examine the movement of immiscible liquids and their vapours in soil
and, (b) to provide data for the verification of numerical transport
models. Work is just beginning on these experiments. The information
presented here provides a brief description of the anticipated research
programme.
Liquid Infiltration

Experiments will be conducted to study the infiltration of
immiscible liquids into sand either partially or fully saturated with
water. A major difficulty with this work is the measurement of the
relative concentration of the wetting and non-wetting phases during
transport through the soil. In situ measurements of electrical (EC) and
thermal (TC) conductivity and dielectric coefficients (DC) have been
proposed. To this point in the research static measurements of DC for
various soil - liquid mixtures have been completed. Figure 12 shows
some results using sand with .different amounts of water, mineral oil

- 434 -

(floater) and carbontetrachloride (sinker). The DC shows good
separation between water and the Innnisclble liquids at equivalent
concentrations .

It should serve as a means for distinguishing between liquids. Toi
a specific liquid, the data in Figure 13 show that capacit.rze (from
which DC is calculated) responds well to changess in concentration.
This property of the measurement will be necessary in order to detect
the dispersed front of an immiscible liquid moving into a water- wetted
soil. In situ measurements of DC have not yet been attempted in this
work but have been done successfully elsewhere (Alharthy and Lange ,
198A). Experiments to quantify the relationships between relative
permeability, capillary pressure and percent liquid saturation are of
major interest and will be attempted in this work. Experimentation
along the lines of that used in the oil industry will be used for
several Immiscible liquids .
Vapour Transport

Experiments will be conducted to trace the movement of vapours in
soil. Soil columns will be exposed to volatile immiscible liquids
either as pools of liquid at the base of the soil or as zones of
unsaturated, liquid-contaminated soils from which vapours can emanate.
Gas chromatography will be used to measure vapour concentrations in the
gas phase and will serve to quantify rates of vapour transport.

- 435 -

SPILL RESPONSE SYSTEM DEVELOPMENT

An important result of this work will be the development of a Spill
Response System (SRS) for innniscible liquids. Although emergency
response models for hazardous wastes do exist, none is specifically and
adequately designed for immiscible liquids spilled onto soil.

It is proposed that the SRS consist of a computer programme which
integrates and interfaces component models currently under development;

- Spill Simulation Model

- Immiscible Liquid Porous Media Flow Model

- Vapour Transport Model

The SRS will function by simulating the sequential behavior of the
liquid including: spill formation, infiltration, transport under
variable saturation, retention in the soil and evaporation. Input
information will be required for liquid properties and data on the spill
conditions, data on soil surface conditions, site hydrogeology. climatic
conditions and local land use conditions. Experience with modelling
fluid behavior in soil has often shown that the accuracy of simulations
is controlled by the quality of the input data for the subsurface soil
conditions. At the time of a spill, data available at a site are likely
to be sparse. Because of this, the SRS will contain default information
that can be used to provide crude estimates of behavior until better
data become available. In many cases, it is expected that certain
conditions such as soil surface slope, liquid volatility, liquid density
and viscosity, depth to groundwater and the presence of coarse soils
will have a dominant effect on the simulation. This will then pre-empt
specific components of the SRS and allow attention to be focussed on the

- 43b -

essential components. This would assist In the design of site
instrumentation. In addition, components of the model can be helpful
long after Che spill occurred, to trace the movement of Immiscible
liquids In soil as an aid to site remedial work. Work on the SRS is in
its Initial stages.

- 437 -

REFERENCES

Alharthi , A. and Lange » J., "Dielectric Properties of Saturated Soils".
Proceedings 2nd International Conference on Groundwater Quality
Research, Tulsa, OK. March 1984.

Arthur D. Little Inc., "S-Area Two Phase Flow Model", Corporate Report.
May 1983.

Bear, J., "Dynamics of Fluids in Porous Media", American Elsevier
Publishing Company, Inc., New York, 1972.

Farquhar , G, and Metcalfe, D, "Gas Migration Modelling". Proceedings of
Symposium on Processes in Landfills . Technlsche Unlversltat
Brauns chweig , Germany , 1982 .

Hillel, D. and Hornberger, G., "Physical Model of the Hydrology of
Sloping Heterogeneous Fields". Soil Sci . Soc. Amer. J. 1979.

Kincaid, D. , "Hydrodynamics of Boarder Irrigation". Ph.D. Thesis .
Colorado State University. Fort Collins, Co. 1970.

MacKay, D, and Matsugu, R. , "Evaporation Rates of Liquid Hydrocarbon
Spills on Land and Water". Can. Jour, of Chem. Eng. Vol. 51. 1973.

Metcalfe, D., "Modelling Gas Transport From Waste Disposal Sites".
M.A.Sc. Thesis , University of Waterloo, Waterloo. 1982.

Osborne, M. , "Numerical Modelling of Immiscible Two-Phase Flow in Porous
Media". M.A.Sc. Thesis , University of Waterloo, Waterloo. 198A.

Sutton, O.G., "Micrometeorology". McGraw-Hill. New York. 1953.

Wall, R., "Numerical Modelling of the Spill of Volatile Toxic Liquids
Over Porous Media". M.A.Sc. Thesis , University of Waterloo,
Waterloo. 1984.

- 438 -

Table 1. Key Words for Che Information Storage
and Retrieval Package (ISRP)

1 .0 Monitoring, Testing, and Detection
■1.1 Vapours In Air

1.2 Detection in Soil

1,2.1 Vapours - In soil

1.3 Groundwater Analysis

1.4 Remote Sensing

1.5 Chromatographic Methods

1.6 Dielectric Coefficient; Conductivities

1.7 Geophysical Methods

1.8 Ultrasonlflcatlon

2.0 Spill Behavior

2.1 Chemical Characteristics

2.2 Evaporation

2.3 Interaction With Hydrogeologic Materials

2.3.1 Migration and Transport

2.3.2 Attenuation

2.3.3 Reactions

3.0 Remedial Work

5.0

4.0

3.1

Liquid Recovery

3.1.1 Pumping

3.1.2 Drainage

3.1,3 Treatment

3.2

In Situ Treatment

3.3

Excavation

3.3.1 Disposal

3,3.2 Restoration

3.4

Encapsulation

3.5

Vapour Control

Mode

4, 1 Transport

4.1,1 Saturated Soil

4.1.2 Unsaturated Soil

Evaporation

4.2.1 From Spills

4.2.2 From Soils
Spilling
Vapours

6.0

4.2

4.3
4.4

Spill Managi

ement

5.1

Chemical Usage

5.2

Managei

nent Models

5,2,1

General Guide
lines

5.3

Legal ]

Restrictions

5.3.1

Government
Regulations

5.4

Safety

5.4.1

Equipment and
Procedures

Field Studies

6.1

Simulated

6.2

Actual

Lab

Studies

- 439 -

Table 2. Assessing The Impact of Hazardous Immiscible Liquids
On Soil and Groundwater: A Literature Review.
Table of Contents ,

1.0 Monitoring, Testing, and Detection

1.1 Vapour Phase

1.1.1 Atmosphere

1.1.2 Subsurface

1.2 Drilling and Sampling
1.2.1 Sample Analysis

1 .3 Chromatography

1.3.1 Theory

1.3.2 Application

1.4 Surface Geophysical Methods
1.4.1 Spill Monitoring

1.5 Aerial Remote Sensing

2.0 Spill Contaminant Migration

2.1 Attenuation

2.1.1 Effects of Organic Liquids on Soils

2.2 Vadose Zone

2.3 Capillary Zone

2.4 Saturated Zone

3.0 Remedial Work

3.1 Hydrodynamlc Control

3.1.1 Migration Control and Contaminant Removal

3.1.2 Treatment Methods for Fluids Extracted

3.2 In Situ Methods for Treatment and Control

3.3 Excavation, Transport and Control

3.4 Encapsulation

3.5 Vapour Control

4.0 Models

4.1 Liquid Transport
4.1.1 Applications

4.2 Vapour Transport

5.0 Spill Management

5.1 Chemical Useage

5.2 Management Models

5.3 Legal Restrictions

5.4 Safety Considerations
5.4,1 Equipment

6.0 Suggested Further Research

- 440 -

650-
600-
550-
500-
450-
400-

Rectangular Spill

Slope- 0.00075 m/m

Flow "2.32 X W-^m^/s per meter

1 r

6 9 12 15 16 21 24 27 30 33 36 39 42
DISTANCE (meters)

Fig. 1- Location of the Spill Front with Time

;Note vertical distortion)

Uniform Flow Depth * 0.50 m

Flow ■ 2.32 X lO"^fn-*/s per meter width
Slope = 0.00075 m/m

200 s \&00 s

Critical Flow Deptti Ao.0084 m

DISJANCe (meters)
Fig.2- Depth Profile for Spill with Time

- 441 -

0.285
0.280
0.270
0.070

, t = Os

d 0.055

> 0.050

0.045-

0.040

0035

0.005-\

Crtticai Flow Velocity • 0.2834 m/s

Flow = 0.00232 m^/s per meter width
Slope « 0.00075 m/m

10 s

400 s

Uniform Flow Velocity 0.0464 m/s

20 25

DISTANCE (m)

Fig.3-VeIocily Profile for Spill

— r-
35

(Note vertical distortion)

Flow= 0.704 -0 004 t m/s
Roughf

10 15 20

RADIAL DISTANCE (meters)

Fig.4-Depth Profile for Radial Spill on to Flat Impervious Surface

Lu(.<iliuM of the hydfjulit )ump

LoCdlioii uf Iho iiinvinK ftntil

r^ r

Flow = 0.704- 0.004 I ni/s
Ruughncss - 0.03 m'A

Clay Loam and Aspha
Sam]

Tnlal sptlk'tl

?0 ?5 20 25 ^Q

RADIUS (meters)

100 )50 2X) 250 JOO

TIME fsec)

Fig.5-Localion of the Hydraulic Jump and Moving Front Fig.6-Accumulated Volume Distribution for Spill on to Clay Loam

- 443 -

570 n

-17 i

560-

>
o

-a

< 550 -\

-170

-167

540-

-164
(ft) (m)

Initidl NAPL disInbuliDii

t' 300-320 days t. 1350 days

— ^ — Arthur D.LittleJnc ^^—
O Osborne a

Soil Surface

20 40

SATURATION (%)

Fig.7-Lithology and NAPL Saturation Profiles (Osborne,1984)

- 444 -

Pumped Ca> veni

t_ Gas Mi^ralion
Monrtorrng Points
(3 probes in each)

100

Fig.8- Monitoring CH^ Migration at the Mississauga Landfill Site

- 445 -

Summarized Soil Strata

\ 1 1

[ 1 1 >
Fine Sand

Loarse bano

r—

and
'3m2

Medium Sand
k:-5 X iC'^m^

Une b
k's' 10"

Coarse Sand
1 1 1 1 1 1 1

I 15.4

>
O

70-

Soil Surface

1 1

;

1 1

10 20 30 40 50

DISTANCE FROM LANDFILL (m)

Fig.9- Methane Migration after Two Days at Mississauga

- 446 -

15.4

Soil Surface

20 30 40 50 bO

DiSJANCE FROM LANDFILL (m)

Fig.lO-Methane Migration after 41 Days at Mississauga

- 447 -

15.4

Soil Surface

20 30 40 50 60

DISTANCE FROM LANDFILL (m)

Fig.11- Methane Migration after 85 Days at Mississauga

- ^148 -

-*^1 10 kHz

w-\

■ Water

A Mineral Oil

• CCI

— t
20

190 -I

SATURATION OF LIQUID IN SAND (%)

Fig.12-Measurement of Dielectric Constant for Various Soil-Liquid Mixtures

100 kHz

180-

'■J

170-\

wo-

Mineral Oil

150

— r —
0.5

1.5

2.0

SATURATION OF LIQUID IN SAND a)

Fig.13-Changes in Capacitance with Increasing Liquid Concentrations in Soi

- 449 -

EFFECTS OF HETALS FROH MINE TAILINGS

ON THE MICROFLORA OF A MARSH

TREATMEirr SYSTEM

PROVINCIAL LOTTERY PROJECT NO. 109

Robert M. Desjardlns and Patricia L. Seyfrled

Departaent of Microbiology, Faculty of Medicine,

University of Toronto

- 450 -

ABSTRACT

The artificial marsh system is an ideal solution for wastewater treatment
in smaller communities. Several artificial marsh systems have been constructed
in Ontario, including Cobalt and Listowel , These two systems possess similar
construction patterns except for the fact that Cobalt's marsh system is built
on a mine tailings basin. Consequently it is necessary to determine if the
metals that may elute from the tailings can exert a serious toxic effect on
the microbial activity required for proper waste treatment.

In this study, the toxic effect of eleven different metals, as well as
metal mixtures, on bacterial isolates from the Cobalt and Listowel marshes
were determined. The agar plate test, the resarurin reduction procedure,
and the ATP luciferin luciferase reaction were used for toxicity testing.
Of the three methods, resazurin reduction was found to be the least effective
in assessing bacterial resistance to heavy metals.

As might be expected, bacteria recovered from the Cobalt marsh were more
metal resistant than isolates from the Listowel marsh treatment system. It
is important to note that strains of eight genera isolated from the marshes
were susceptible to lower concentrations of metals when metal mixtures were
tested.

Sampling sites in the marsh systems were monitored for heterotrophic
bacteria, total col i forms, fecal coliforms, Escherichia col i , fecal streptococci,
^""^ Pseudomonas aeruginosa . Results showed that both the Cobalt and the Listowel
artificial marsh treatment systems were responsible for substantial reductions
in levels of these bacterial parameters. By comparison, the natural marsh
system at Cobalt was much less efficient at reducing bacterial numbers from
the inflow to the outflow sites.

- 451 -

INTRODUCTION

Ecological processes such as the microbial decomposition of plant and
animal litter, critically affects the quality of aquatic and terrestrial eco-
systems. Damage to these ecological processes may have greater environmental
consequences to the ecosphere, than damage to a particular plant or animal
species.

Because microbes are not visible to the naked eye, and because the ecological
processes under their control are subtle and overt, they tend to be overlooked.
Nonetheless, microbes are sensitive to pollution, and Inhibitions of microbial
activity Is accompanied by reduction In the ecological processes under their
control .

Microbial invisible Injury may be illustrated by the example of mineralization
of plant and animal litter, which Is an important nutrient regeneration process
needed to maintain fertility of aquatic and terrestrial ecosystems. The mineral-
ization process by microbial decomposers releases inorganic carbon, nitrogen,
sulfur; phosphorus and other chemicals that are essential for, and are assimilated
by the indigenous photoblota. An adverse effect on microbial mineralization
by heavy metals for example, decreases the ability of that ecosystem to support
an abundant flora and fauna.

To date, researchers have confined their work to studies of the effect
of a single chemical on one particular species. As Davis (1) has pointed
out, although we are seldom exposed to only one specific chemical, chemical
synergies have rarely been considered in environmental risk assessments.

In this project, we plan to examine the effect of 11 different metals
as well as mixtures of the metals, on the heterotrophic bacteria In a marsh
treatment system. The artlfical marsh selected for the study is unique In

- 4r,:> -

that it is situated in a mine tailings basin and thus may be affected by the
heavy metals eluted from the tailings. Singleton and Guthrie (2) have shown,
for example, that minimal levels of metals, such as 40 ppb, are sufficient
to decrease the bacterial diversity of an ecosystem. The pH of the system
may also be an important factor since Babitch and Stotzky and others (3,4)
have shown that the toxicity of the metal increases as the pH increases.
The overall objectives of the study were as follows:

1) To assess the effect of metals on the normal microflora of a marsh
treatment system;

2) To compare the metal susceptibility of heterotrophic bacteria isolated
from a mine tailings basin marsh located in Cobalt and from a normal marsh
system located in Listowel; and

3) To determine the efficiency of the above marsh treatment systems,
with respect to reduction of micro-organisms of fecal origin.

- 453 -

fCTHODS

Sampling Sites
COBALT:

Water and soil samples were collected from the two artificial and one
natural marsh treatment systems, located in Cobalt, Ontario. Connecting streams
and a nearby pond (Figure 1) were also sampled. All three marsh treatment
systems received an input of raw sewage.
LISTOWEL:

Systems 4 and 5 were selected for sample analysis because they most closely
resembled those constructed in Cobalt. Both systems received effluent from
an aeration cell (pretreatment) . Water samples were also taken from the sewage
distribution centre, treated water exit, and west cell (FIGURE 2).

PROCEDURES

Sample Collection :

Surface water, sewage and effluent samples were collected in sterile,
sodium thiosulphate treated bottles, and chilled during transport to the labor-
atory. The pH and temperature of the water samples were determined on site
at the time of collection. Samples were obtained from the Cobalt marsh in
November, 1983 and in June and August, 1984; the Listowel marsh was sampled
in May and August, 1984.
Bacterial Analysis :

Appropriate dilutions of each sample were made, and 0.1 mL of each sample
was spread In triplicate on casein-peptone-starch agar (CPS) (5), standard
plate count agar (SPC) (Difco), and heterotrophic plate count agar (HPC) (6).

- 454 -

One set of plates was incubated at 21*^C for 7 days and another set at 3iOC
for 2 days after which counts of the colonies were made to determine the number
of aerobic heterotrophic bacteria present in each sample.

For further heterotrophic bacterial identification, all the colonies
on a plate (or sector of a plate) were selected in order to obtain a representative
of the total bacterial population present. The colonies were picked and streaked
three times on CPS agar for purity.

Approximately 200 organisms were isolated from the Cobalt Marsh treatment
system at each sampling event in November and June and identified to the genus
level using standard methods. One hundred and twenty organisms were collected
from the Listowel marsh at each sampling period in May and August and identified.

Analyses of the samples for total coliforms, fecal coliforms, Escherichia
coll, fecal streptococci, and Pseudomonas aeruginosa were performed according
to Ontario Ministry of the Environment specifications,
rCTAL SOLUTIOMS

Graded concentrations of Pb(N03)2, FeCla, Al CI3.6H2O, HgCl2s (Fisher)
CUCI2.2H2O, CrCl3.6H20, CoCl2.6H20, ZnS04.7H20 (Sigma Chemicals) Na2HAs04.7H20,
NiCl2.6H20 and CdCl2.2 1/2 H2O (BDH chemicals) were made using sterile deionized
distilled water. Metal solutions were prepared fresh each day.
AGAR PLATE TOXICITY TEST (APTT)

This procedure was described by Lui and Kwasniewska (7) for the rapid
assessment of chemical inhibition to microbial populations. Bacterial isolates
obtained from the Cobalt and Listowel marshes were examined for their response
to the following concentrations of heavy metals: 10,000, 5,000 4,000, 2,500,
1,000, 500, 200 and 100 ug/mL for Al , As, Cd, Co. Cu, Fe, Ni , Pb. and Zn and

- 455 -

100. 50, 25. 10, 1, 0.5 and 0.1 ugM for Hg. Mixtures of the metals used
are listed in Tables 5 and 6. The toxicity test procedure was repeated three
times usi ng 30 pure cul tures from each sampl i ng event i n November and June
at Cobalt, and 20 cultures from each of the May and August samplings at Listowel,
RESAZURIN REDUCTION

This procedure has been described by Lui et al (8,9). The incubation
time of 30 min, suggested by Lui (8) was extended to up to 3 hours or until
significant reduction of the dye had occurred. Metal solutions were prepared
so that the final concentrations of the metals were 75, 50, 25 and 10 ug/mL
for Al, Fe, Ni , Zn, Cd, and Co; 50, 20, 10, and 5 ug/mL for Cu; 5000, 1000,
500 and 100 ug/mL for As; and 5, 1, 0.5 and 0.1 ug/mL for Hg. Metal mixture
concentrations used were the same as listed in Tables 5 and 6.

Six pure cultures of Bacillus, Klebsiella , Enterobacter , Aeromonas , Pseud-
omonas and Escherichia spp., collected from the Cobalt marsh during the November
sampling period, were chosen for use in this experiment.
ATP BIOASSAY

The firefly luc if erase assay of intracellular bacterial adenosine triphosphate
was developed in our laboratory to measure the toxic effects of metal ions
on aquatic microorganisms (4). This procedure was further modified by spectro-
photometrically adjusting the cell suspension to 0,5 O.D. { A 625 nm) before
inoculation of the fresh broth, to obtain log phase cells. Ten mL of nutrient
broth was substituted for 10 ml of minimal media broth for optimal growth
of the Bacillus isolate. To date, only cadmium has been tested, at levels
of 100, 50, 25, 10 ug/mL, on the same isolates used in the resazurin reduction
experiment.

- 456 -

RESULTS

Of the three marsh treatment systems in Cobalt (Figure 1) only artificial
system #1 and the natural swamp were functional. Due to exfiltration problems
in artificial marsh system #2, analysis of this site did not extend beyond
the first sampling period.

Table 1 shows that the heterotrophic plate counts from the Cobalt marsh
treatment system sites ranged from 6,0 x 10^ to 9.5 x 10^ bacteria per mL.
The counts at most of the sampling sites were found to vary by a factor of
10 between summer and fall samples. The percent distribution of aerobic hetero-
trophs from the Cobal t and Li stowel marshes at di fferent sampl ing times i s
presented in Table 2. Acinetobacter and Flavobacterium were the two most
commonly identified genera. Those cultures that were not identified are undergoing
further studies.

Table 3 shows that both Cobalt marsh treatment systems were effective
in reducing the levels of all the bacterial parameters tested from the inflow
sites (2 and 7) to the outflow sites {5 and 8) indicated in Fig. 1. The Listowel
marsh system was also responsible for a drop in numbers of heterotrophic bacteria,
total coliforms and fecal streptococci from site 1 to site 3(Fig. 2). Levels
of the other parameters were not high enough to measure with the dilutions
used. The pond at Cobalt {site 6) which had no municipal sewage input had
a compara ti vel y 1 ow number of col i form bacteri a ( 40 per 100 mL ) . However ,
the receiving stream (site 9) was found to have high counts of fecal coliforms
(1,8 X 105 per 100 mL).

A comparison of the different media used for the isolation and enumeration
of the heterotrophic bacterial population is presented in Table 4. The suggested
incubation temperature for CPS is 21°C, and for HPC and SPC, 35°C. When the

- 457 -

recommended incubation was followed, the counts were observed to vary from
3.1 X 10^ and 8.6 x 10^ to 1.4 x 10^ CFU per mL (on SPC. HPC, and CPS respectively)
for the same water sample taken from site 3.

A minimum of two strains of each of the genera listed in Tables 5 and
6 were subjected to metal toxicity testing using the agar plate method. The
metal sensitivity patterns of the organisms are given in the tables. It was
found that all of the organisms of the same genera isolated from Cobalt in
November exhibited essentially the same metal resistance pattern. The organisms
tested tended to be sensitive to much lower concentrations of the metals when
the metals were present in mixtures. For example, the Bacillus isolate was
sensitive to 500 ug/mL of Al when tested singly but was sensitive to 20 ug/mL
when tested in combination with other metals,

Thi rty i sol ates, obta i ned from Cobal t i n the June sampl i ng peri od and
identified as members of the six genera listed in Table 5, were also subjected
to metal toxicity testing using the agar plate method. These strains were
found to have metal sensitivity patterns similar to the November isolates
previously described.

The twenty cul tures isolated from Lis towel and classified as members
of the genera Proteus , Flavobacterium and Chromobacter showed a wide range
of metal sensitivity patterns, even within the specific genus level (Table
6). Most of the Listowel cultures were sensitive to ^ 100 ug/mL Cr and
^lug/mL of Hg.

The resazurin reduction method was used to determine the concentration
which inhibits 50% of the organisms' dehydrogenase enzymes (IC50). The results
were plotted and the IC50 determined (Figure 3). The sensitivity patterns
to seven metals and metal solutions are presented In Table 7.

- 458 -

A comparison of metal sensitivity patterns obtained using the APTT and
resazurin reduction experiments Is shown in Table 8. In most cases each testing
procedure displayed its own unique sensitivity pattern.

Figures 4 to 9 illustrate the toxic effects of cadmium on six different
genera of bacteria isolated from Cobalt. Triplicate experimental trials were
performed on Enterobacter and similar results were obtained each time. Each
of the different genera tested displayed Its own sensitivity pattern to cadmium.
DISCUSSION

The number of aerobic heterotrophs isolated at different sampling times
varied by a factor of 10 between fall and summer samples (Table 1). The higher
counts that were obtained In the summer might be expected since the temperature
of the water from which the isolates were obtained was very close to the incubation
temperature used for plate counts. Trentham and James (10) have shown that
temperature Influences the bacterial dynamics of a coimiunlty. The mesophlllc
bacteria predominate In the summer months, and usually do not grow in winter
months (hence lower counts). On the other hand, the psychrophilic organims
are not able to survive the higher temperatures of the summer.

In this study, the incubation temperature of 21^0 would not be suitable
for isolation and enumeration of the psychrophilic organisms in the November
samples, and thus lower counts were observed. Lower Incubation temperatures
win be compared in future studies.

The frequency of distribution of each genera of heterotrophic bacteria
did not differ to any great extent for the different sampling times, at either
Cobalt or Listowel. The specific sampling sites chosen showed only a slight

- 459 -

variation In the types of heterotrophs identified. For example, those sampling
sites close to the areas of sewage Inflow had a higher percentage of fecal
coll form bacteria.

Acinetobacter and Flavobacterlum were the most abundant genera Isolated
at both Cobalt and Listowel . It Is Interesting to note that Acinetobacter
Is often found In conjunction with high organic loading (11). The data presented
herein support this fact. Members of the genus Flavobacterlum are also common
inhabitants of soil and water.

The efficiency of each type of marsh treatment system was assessed using
standard bacterial parameters (Table 3). The artificial system #1 (sampling
sites 2,3,4 and 5) at Cobalt proved to be effective in reducing the number
of fecal conforms exiting from the system (FC counts at the sewage entrance
were calculated to be 9.4 x 10^ per 100 mL and were reduced to less than 100
FC per 100 mL at the treated sewage exit). The receiving stream, into which
the marsh treatment system emptied, was Itself grossly polluted with fecal
conform counts of 2.1 x 10^ per 100 mL at site 1, upstream from the marsh
system. The counts were somewhat reduced, to 2.1 x 10^, at a downstream sampling
site 10. This reduction in counts, from the upstream to downstream sites,
would indicate that the outflow from the marsh treatment system was not overloading
the receiving stream with respect to coll form organisms.

The natural marsh system at Cobalt (sampling sites 7 and 8) was not as
effective In fecal coliform reduction; counts of 8.2 x 10^ per 100 mL were
only reduced to 1.1 x 10^ per 100 mL. Samples taken further downstream, at
site 9, were found to contain the same magnitude of fecal conforms as the
marsh outflow site.

In comparison, the counts of bacterial parameters, taken from the Listowel

- 460 -

Marsh and listed in Table 3, indicated that the aeration pretreatment was
very effective in reducing fecal col iform numbers, since few were detectea.
The data suggest that the two artificial marsh treatment systems are very
efficient at reduction of fecal coliform bacteria, as compared to that of
the natural marsh treatment system located in Cobalt, Containment of Cobalt's
natural marsh system to resemble system 5 at Listowel, should improve the
ability of the former system to reduce the fecal origin bacteria exiting the
system.

Although there is no appropriate medium for the isolation and enumeration
of all types of bacteria, we compared three different types of media to see
which medium would be most suited for the isolation of aerobic heterotrophs.
The plated samples were incubated at both 35° and 21°C. Table 4 demonstrates
the different counts of bacteria using different media and at different temper-
atures. At 350c. HPC > SPC >CPS and at 21°C, CPS > HPC > SPC in enumeration
of the heterotrophs. The counts taken at Zl^C in most cases were higher than
those obtained at 350C.

The constituents of HPC and CPS are quite similar, and this was reflected
in the similar counts that were produced. SPC is a high nutrient medium and
many environmental strains may not be able to tolerate this type of nutrient
shock.

When each medium was incubated according to the prescribed requirements
and its ability to recover heterotrophs was compared, CPS was determined to
be a superior to HPC or SPC (counts were 1.4 x 10^, 8.6 x 10^, and 3.1 x 10^
CPU per mL respectively). Temperature may have influenced the counts since
35OC may have been a greater selection factor than 21°C. Further analysis
of our 21°C isolates has shown that less than 50% were able to grow at 35°C.

- 461 -

The APTT was found to be very useful in screening the relative heavy
metal sensitivities of the different genera of bacteria isolated from Listowel
and Cobalt. Those isolated from Cobalt (Table 5) tended to be more resistant
to chromium, copper and cobalt, than those isolated from Listowel (Table 6).

Intragenera differences with respect to heavy metal sensitivities were
noted between isolates taken from Listowel , but not to any great extent in
those bacteria isolated from Cobalt. Perhaps there is a selective pressure
due to the presence of the mine tailings (metal elution) in Cobalt, that is
not present at Listowel. Cultures taken later in the year from each marsh
treatment system followed the same type of heavy metal sensitivity pattern
noted previously.

The APTT experiment was performed in triplicate and results were always
consistant. In most of the trials, if the metal sensitivity concentration
was different, it only differed by one level above or below the other trials.
Results were easy to read and interpret because the presence or absence of
growth was quite evident.

The resazurin reduction procedure was deemed not acceptable for assessing
relative heavy metal toxicities. The results in Table 7 depict those ICsq's
that we were able to calculate.

After the metal solutions were added to the resazurin reduction broth,
a precipitate developed. The only metals that did not precipitate with the
broth were mercury and arsenic. Low concentrations of metals ( ^ 40 ug/mL)
did not precipitate with the broth; however, no apparent differences in dye
reduction were noted at concentrations lower than 40 ug/mL (except for copper
and mercury) .

- 462 -

f^igure 3 displays the IC50 curves obtained with arsenic, Pseudomonas
had more dye reduction occurring in the presence of arsenic than the control.
This may be explained by the Arndt-Schultz effect (12) whereby low concentrations
of metal may enter the plasma membrane which becomes distorted. An Influx
of nutrients and resazurin may enter the cell, stin.ulating bacterial growth,
leading to increased reduction of resazurin.

There appeared to be a threshold effect with respect to arsenic toxicity,
i ,e. there was a level of metal which the bacteria could tolerate before inhibition
of the dehydrogenase was evident (Figure 3).

This experiment was not useful in assessing the relative toxicities of
Al , Cr, Fe, Pb and Zn, as no enzymic inhibition was measured at the levels
tested. Repeat experiments performed with higher concentrations of metals
lead to inaccurate results because some of the metals were able to reduce
the resazurin in the absence of the bacterial Inoculum.

One further problem with this procedure is that none of the isolates
were able to reduce the resazurin at the same rate. Incubation times varied
up to three hours before an appreciable dye reduction had occurred. Use of
only one type of organism would give no indication of phenomena which may
actually occur In the environment. Further, this biased sampling does not
take into account possible resistance mechanisms of the test microbe.

Table 8 displays the different heavy metal sensitivity patterns as determined
using the APTT and IC50. None of the isolates had the same resistance pattern
of most toxic to least toxic metals. It is only possible to state that As
was the least toxic while Hg was the most toxic metal.

The resazurin reduction test is a short term assay, where the experiment
is usually run for less than one hour. In comparison, the APTT is an all

- 463 -

or none type of experiment where the toxicant remains I'n contact with the
bacterial cells for 24 hours. The resarurin reduction procedure fails to
take into account the possibility that the bacterial cell may be able to accom-
modate to stress caused by the toxicant, but this may take some time after
the time limit set by the experiment. In this type of short term exposure,
it is not possible to measure the effects of recovery.

The bacterial cultures used in the resarurin reduction experiment were
also subjected to cadmium, and the effects measured by the intracellular ATP
bioassay. Figures 4 to 9 display the toxic effects of cadmium on the six
different genera.

The Pseudomonas isolate, as shown in Fig. 4, appeared to be relatively
resistant to less than 50 ug/mL of cadmium. Initially, the culture appeared
to be stressed to 50 and 100 ug/mL, but 6 hours after the addition of 50 ug/mL
of cadmium, recovery became evident. The degree of toxicity was proportional
to the concentration of cadmium, with the most significant effects at 100
ug/mL. The Enterobacter (Figure 5) and Klebsiella (Fig. 6) display essentially
the same pattern as the Pseudomonas isolate.

Figures 7 to 9 illustrate the toxic effect that cadmium has on the Aeromonas ,
Bacillus and Escherichia isolates. All were initially stressed to all levels
of cadmium, Aeromonas and Bacil lus were able to recover to an appreciable
extent to 10 ug/mL cadmium, but not to the other levels tested. The Escherichia
isolate was not able to recover from the toxic effects of cadmium (Fig. 9).
There was a short period of recovery at about 3 hours, followed by a sharp
decline in amount of intracellular ATP detected.

The intracellular ATP assay is a sensitive, accurate means of determining
the toxic effects of metals on the bacterial cell. This bioassay, unlike

- 464 -

the resazurin reduction procedure, has shown that some of the isolates are
able to recover from the various concentrations of metal solutions added.

SIffMARY

The artificial marsh treatment systems located at Cobalt and Listowel
were more efficient at reduction of fecal col i forms than the natural marsh
system in Cobalt. The bacteria isolated from the Cobalt marsh system were
found to be less sensitive to heavy metals, and the level of sensitivity appeared
to be uniform in each specific genera tested. Most Listowel isolates were
sensitive to less than 100 ug/mL of Cr, and displayed intrageneric differencrs
with respect to the other heavy metals tested. The resazurin reduction procedure
was not useful in determining heavy metal resistance patterns. On the other
hand, the intracellular ATP analysis proved to be an efficient accurate, repro-
ducible method of determining the toxic effect of metals on the bacterial
cell.

- 465 -

REFERENCES

1. Davis, D.L. 1979. Multiple risk assessment: preventive strategy
for public health. Toxic Subst. J. U 205-225.

2. Singleton, F.L. and R.K. Guthrie. 1977. Aquatic bacterial populations
and heavy metals - 1. Composition of aquatic bacteria in the presence
of copper and mercury salts. Water Res. 11: 639-642.

3. Babich, H. and G. Stotzky. 1977. Sensitivity of various bacteria,
including actinomycetes, and fungi to cadmium and the influence of pH
on sensitivity. Appl . Environ. Microbiol. 33_' 681-695.

4. Seyfried, P.L. and C.B.L. Morgan. 1983. Effect of cadmium on lake
water bacteria as determined by the luciferase assay of adenosine
triphosphate. Aquatic Toxicology and Hazard Assessment: Sixth
Symposium. ASTM STP 802, W.E. Bishop, R.D. Cardwell, and B.B. Heidolph,
Eds., American Society for Testing and Materials, Philadelphia,

pp. 425-441.

5. Staples, D.G. ^ r.d J.C. Fry. 1973. A medium for counting aquatic hetero-
trophic bacteria in polluted and unpolluted water. Jour. Appl.

Bact. 36: 179-181.

6. ASTM Working Document. May, 1983. Proposed test procedure: Heterotrophic
bacteria in water.

7. Lui, D. and K. Kwasniewska. 1981. An improved agar plate method for
rapid assessment of chemical inhibition to microbial populations.
Bull. Environ. Contam. Toxicol. 27: 289-294.

- 466 -

8. Lui, D. and W.M.J. Strachan. 1979. Characterization of microbial
activity in sediment by resazurin reduction. Archiv f. Hydrobiologie,
Beih. 12: 24-31.

9. Lui, D., K. Thompson, and K.L.E. Kaiser. 1982. Quantitative structure -
toxicity relationship of halogenated phenols or. bacteria. Bull.
Environ. Contam. Toxicol. ^: 130-136.

10. Trentham, J.N. and T.R. James. 1981. Seasonal selection in a freshwater
heterotrophic bacterial community. Microb. Ecol . 7_: 323-330.

11. Seyfried, P.L. 1973. Sampling bacteria in Lake Ontario and the
Toronto Harbour. Proc. 16th Conf. Great Lakes Res., Internat. Assoc.
Great Lakes Res. pp. 163-182.

12. Doyle, J.J., R.T. Marshall, and W.H. Pfander. 1975, Effects of cadmium
on the growth and uptake of cadmium by microorganisms. Appl. Microbiol.
29; 562-564.

- 467 -

Table 1. SEASONAL DIFFERENCES IN THE AEROBIC HETEROTROPHIC

BACTERIAL POPULATION ISOLATED FROM THE COBALT MARSH TREATMENT SYSTEM

Site ^

November Sample

June Sample

August Sample

6.6 X 10^

1.9 X 10^

4 X 10^ ^

6.2 X 10^

1.5 X 10^

3.8 X 10^

1.4 X 10^

2.3 X 10^

7.3 X 10^

1.9 X 10^

6 X 10^

2.0 X 10^

4.2 X 10^

6.5 X 10^

7 X 10^

1.3 X 10^

7
B

9.5 X 10^
1.9 X 10^

7 X 10^
1.6 X 10^

1.6 X 10^
3.9 X 10^

2.4 X 10^

1.5 X 10^

4.7 X 10^

1.9 X 10^

1.4 X 10^

4.8 X 10^

Water

temperature

range

2° - 11°C

12 - 16°C

17 - 19°

pH range

6.5 - 7.3

6.5 - 7.5

6.4 - 7.4

see Fig. 1

CFU per mL

- 468 -

Table 2. PERCENT DISTRIBUTION OF IDENTIFIED GENERA FROM COBALT

AND LISTOWEL MARSHES

Cobalt

Listowel

Organism

November

June

May

August

Acinetobacter

Flavobacterium

Corynebacterium

Pseudomonas

Klebsiella

„

Streptococcus

Staphylococcus

Pasteurella

Escherichia

Bacillus

Alcal igenes

Chromobacterium

Enterobacter

Proteus

Aeromonas

Salmonella

Unidentified

Month samples were collected.

TABLE 3.

EFFICIENCY OF TREATMENT SYSTEMS IN COBALT, BASED ON BACTERIAL PARAMETERS

Cobalt

Lis towel

Sampling Site

Heterotrophs per mL

PS'

1 Creek

1.9 X

10^

6.8 X 10^

2.1 x.lO^

2.4 X 10^

' 8.1 X

10^

1 3
9.2 X IQ-^

2 Sewage inflow

1.5 X

10^
10^

3.6 X 10^

9.4 X 10^

4.7 X 10^

6.5 X

10^

1.5 X 10^

3 Near sewage exit

1.9 X

4.0 X 10^

<100

2.0 X

10^

<100

5 Treated water exit

4.1 X

10^

5.0 X 10^

<100

1.5 X

10^

<100

6 Natural pond

1.3 X

10^

3 X 10^

4.0 X 10^

3.7 X

10^

<100

7 Sewage entrance

1.58 X

10^
10^

1.3 X 10^

8.2 X 10^

3.9 X

10^

2.6 X 10^

8 Treated water exit

3.9 X

1.8 X 10^

1.1 X 10^

9,2 X 10^

7.5 X

10^

5.7 X 10-^

9 Receiving stream

4.7 X

10^

1.8 X 10^

1.6 X 10^

9.1 X

10^

8.1 X 10^

1 Sewage entrance

1.2 X

10^

1.2 X 10^

<1G0

<100

1.8 X

10^

<100

2 Midway in trench

10^

system

3.2 X

2.9 X 10^

<100

1.8 X

10^

<100

3 Sewage exit

6 X

10^

1.0 X 10^ ,

<100

1.1 X

io3

<100

6 West cell

3.1 X

10^

9 X 10^

<100

1 X

102

<100

.'1'

Bacterial parameters:

heterotrophic bacteria

TC = total coliforms

FC = fecal coliforms

EC = Escherichia coli

FS = fecal streptococci

PS = Pseudomonas aeruginosa

Counts per 100 mL of sample

- 470 -

Table 4. COMPARISON OF DIFFERENT MEDIA FOR ISOLATION AND ENUMERATION
OF AEROBIC HETEROTROPHS FROM THE COBALT MARSH TREATMENT SYSTEM

35"

Incubation Temp.

21"

Incubation

TeiTip .

Sample Site

HPC^

CPS^

SPC^

HPC^

CPS^

SPC

3
4

8.6 X 10^
3.4 X 10-^

2.4 X 10^
1.4 X 10^

3.1 X 10^
6.3 X lO'^

2.03 X 10^
6.5 X 10^

1.4 X 10^
7.3 X 10^

5.1 X 10^

4.2 X 10^

see Fig. 1

HPC = heterotrophic plate count agar
CPS = casein - peptone - starch agar
SPC = standard plate count agar

Table 5,

Sensitivity of Bacteria! Isolates obtained from Cobalt to Single Concentrations of Metals
or to Metal Mixtures.

Organism

Bacillus

500*

Pseudomonas 2500

Ccrynebacterium i 4000

Aeromonas ' 500

Flavobacterium

Escherichia ; 500

1000
500
500

1000

Heavy Metal Sensitivity Pattern (ug/ml)

2500

500

2500 i 1000 I 500

200 2500

500

500 I 200

1000 4000

1000

2500 i 1000 I 2500

<100 ; 2500 I 10^ i 4000

>10*

>10'

1000 200 I 500 >10

10^ i 10^1 >io^

200 ; 200 ' 200 ' 2500 ' 200 ■ 500 ! >10^

<1
<1
<1

Mixture
12 3 4 5 6

10'

500

5000 I 4000

10^ '• 4000

S R S R S S

S S S S R S

R R R R S S

1000

500

2500

1000

; 500

Mixture 1.
2.
3.
4.
5.
6.

200 ug/ml of Co Al Zn Cr Cu

100 ug/ml of Co Al Zn Cr Cu

200 ug/ml of Fe Al Co Cr Ni

100 ug/ml of Fe Al Co Cr Ni

10 ug/ml of Fe As Al Zn Cd Cr Cu Co Ni + 1 ug/ml Hg

20 ug/ml of Fe As Al Zn Cd Cr Cu Co Ni + 1 ug/ml Hg

Values represent the mean of three tests for each strain.

Table 6.

Sensitivity of Bacterial Isolates obtained from Listowel, to Sinale Concentrations of Metals
or to Metal Mixtures

Organism

Proteus

Flavobacterium 3M

Chromobacter 3E

5P 1 500*

SB ' 2500

4C ; 500

5iR 500

1000
2500
1000

<100

2500

<100

500

1000
250
500

<100

500

1000

500

<100

200

Heavy Metal Sensitivity Pattern

4000
500
500
500

1000

200

4000

<100

500

200

4000

200

1000

500

1000

<100

1000
2500
2500
10^

10^

2500

500

500
1000
<100
2500

2500
4000
1000

500

>10'

>10

>10*

>10

>10'

.5
.5

10
.5
1

<100

2500

200

5000

<100
<100
2500

<100

Mixture
1 2 3 4 5"^ 6"*"

5000 S S S S S S

lo'^ S S S S R R

5000 S S S S S S

2500 i S S S S R R

10^ ! S R S R R R

10^ S S S S S S

2500 S S S S R R '

I -J

2500 ; S S S S R S •

+ has .1 pg/mL Hg instead of 1 yg as in Table 5.

Values represent the mean of three tests for each strain,

- 473 -

Table 7. IC^p Values Calculated using the Resazurin Reduction Method

^^50

(yg/mL)

Culture

Metal Mixture

Bacillus

600

140

11000

Klebsiella

100

9.5

110

5000

Enterobacter

5000

Aeromonas

230

6000

Pseudomonas

400

1.2

15000

3.8

Escherichia

6000

3.4

Stock solution contains 10 parts of each of Al, Co, Cr, Cd, Cu, Zn. Ni ,
As and Fe plus 1 part of Hg

Insufficient data to obtain IC

- 474 -

Table 8. COMPARISON OF METAL SENSITIVITY PATTERNS USING THE AGAR PLATE
TOXICITY TEST (APTT) AND THE RESAZURIN REDUCTION PROCEDURE FOR

DETERMINATION OF IC^q

Metal
Culture Co Cu Cd Hg As Ni

Bacillus APTT

Pseudomonas APTT

Escherichia APTT

Enterobacter APTT

IC5Q 30 9 - .8 5,000

Aeromonas APTT KOOO 1.000 ?no <5 >10^ 500
IC5Q 230 17 - .6 6000

1,000

500

200

600

100

500

200

400

200

2,500

200

500

2,500

100

1,000

200

230

>io4

1000

11,000

140

>10^

2,500

1.2

7000

4.8

10^

500

6000

>10^

2,500

Natural Pond

// //// /// /'/// / ////■'././ i.i/ijjj/ijjj mmt

Mgure 1:

1-10 Sampling Sites at Cobalt

Artificial Lagoon #1

Artificial LagoQn I>1

Treated Hater Outflow

System IV

System V

■si

Aerated Sewage Entrance

Figure 2j Sampling sites 1-6, Listowel

West Cell

- 477 -

-0.9

-1 .2

Figure 3; Effect of Arsenic on Selected Isolates Based on
Inhibition of Dehydrogenase Enzymes (IC ).

Legend:

. . Pseudomonas

X— .— X Aeromonas

o o Klebsiella

- 478 -

Fip.ure ^4 : Effect of Cadmium ions on Fseudomonas as
measured by the lucifcrase assay.

]H; ATI'/I.

- n7<) -

I'iciirf S; i:M(mI ol Cadmium on l intorobactf r ns inonisiirod tjy rirofly
liiciforasp assay (intracellular ATP ur/iiiI).

f) iif*/m"L^

10 u

25 iift/lTii

r. '\Ti7i.

Timo (hours)

100 nc/nil

- 480 -

uc A'rr/i.

Figure 6: Toxic effect of cadmium on Klebsiella as
indicated by intracellular ATP levels.

'I'liric ( hours)

- 481 -

rifiiiro 7 : Toxicity of cadmium lor an Aeromonas sp. isolate
analysed using the firefly luciferase assay
(intracellular ATP fJg/L),

If. All'/I.

'I'i niP ( hoiU'.s )

- 482 -

tic. ATI'/ 1.

rif^uro 8 : KfTnct of cadmium on a sfwane isolnto oT

liari 1 1 us as Tnoa.surnd by thn iiit rar(*l 1 ul nr ATI''

Time (hours)

f^'t-.^'^fs^-^ .-V :* ■■.■•■i¥.^t ^■H^f'V ■:y::^-v:: . f' ■;

'■' ''■^■/■^- ?::';

2+
FiRure 9: Kffoct of Cd on rscherichia as measured by the

Intracellular ATP Assay.

'I'iiiio (hours)

- 485 -

REVISED MONITORING SCHEIE FOR
PERSISTENT AND TOXIC ORGANICS IN
GREAT LAKES SPORTS FISH

J. A. Coburn and H. Huneault

ZENON ENVIRONMENTAL INC.

845 Harrington Court

Burlington, Ontario

and

Gerald Rees and George Crawford

ONTARIO MINISTRY OF THE ENVIROIMENT

- 486 -
ABSTRACT

Protocols for the analysis of a broad range of
synthetic organic compounds have been Identified and evaluated on
fortified and unfortified fish tissue samples. Recoveries of the
trace organlcs have been determined at low parts per million and
parts per billion concentrations. Sample extracts were processed
using gel permeation chromatography (GPC) for elimination of
lipids from the trace organic fraction prior to full scan gas
chromatographic-mass spectrometric (GC/MS) analysis. Data will
be presented on the sample preparation protocol and the GC/MS
analyses of fish samples from the Great Lakes. These latter data
will Include contaminant identifications as well as spatial
trends for selected chemicals In several fish species.

- 487 -

Over the past two decades the number of detections of
trace organic contaminants in fish samples has been Increasing
considerably. These analysis have detected the presence of PCB,
DDT and metabolites, chlordane, benzene hexachloride isomers
including lindane, dieldln, endrin, chlorinated benzenes, mirex,
mirex metabolites, and chlorinated styrenes (COA Report, 1981).
The International Joint Commission has reported that in addition
to the usual chlorinated hydrocarbons, phthallc acid esters,
volatiles such as chloroform, bromoform and tetrachloroethylene,
Toxaphene, polyaromatic hydrocarbons (PAH) and alkylated and
chlorinated PAH and a broad array of aliphatic hydrocarbons have
been Identified in fish.

Numerous authors have more recently reported the
presence of 2,3,7,8-tetrachlordibenzo-p-d1oxin at part pere
trillion concentrations (Kuehl, 1981; Ryan, 1982; Harless, 1980).
Additional analysis have also revealed the presence of ultra-
trace (part per trillion) levels of other chlorinated dioxins and
polychlorlnated-dibenzofurans (Kuehl, 1981).

Kuehl, (1980) has also reported the presence of
hexachloro and heptachloro styrenes, pentachlorophenol and
pentachl orobenzyl alcohol in fish samples in the Great Lakes
watershed.

Chlorinated toluenes and chlorofluorotoluenes have been
reported in fish tissue samples collected from the Niagara River
(Yurawecz, 1979). Polychlorinated diphenyl ethers have been
reported in sediments and fish collected from Whitby Harbour in
Lake Ontario (Coburn 1981) with levels higher than the PCB.

- 488 -

Polychlorlnated naphthalenes have been detected and
identified In fish from the Titabawassee River which flows into
Saginaw Bay of Lake Huron (Kuehl, 1980) at very low parts per
billion concentrations.

Hexachlorobutadiene and hexachlorocyclopentadiene have
been reported to be in numerous hazardous waste landfills along
the Niagara River in the New York State (Intereagency Task Force
on Hazardous Wastes, 1979) however only the hexachlorobutadiene
was detected in the fish samples from the Niagara River

(Yurawecz, 1979).

A comprehensive analysis of water and sediments
adjacent to hazardous waste landfills in Niagara Falls, N.Y.
(Elder, 1981) has detected chlorobenzenes, chlorotoluenes,
polyaromatic hydrocarbons, PCB, chlorophenol s, fluorinated
aromatics. and a series of unchlorinated and chlorinated benzyl
derivatives. There have been no reports in the literature of the
detection of this latter group of organics in fish samples from
Niagara River or Lake Ontario.

The report of Hesselberg, (1982) also identified two
halogenated contaminants found in Great Lakes fish samples which
were not previously identified as pollutants of concern by the
International Joint Commission. They were 3-chloro-l-propynyl-
cyclohexane and 3-bromomethyl-cyclohexene with the latter
compound being detected in all Great Lakes samples tested.

- 489 -

These reports reveal that as analytical techniques
continue to Improve, there are Increasing detections of trace
organic chemicals In the fish of the Great Lakes at ever
decreasing concentrations.

There have been basically four main techniques employed In
the extraction of persistent and toxic organlcs such as DDT,
chlordane, PCS, mirex, etc. from biological tissues. These are:
i) mechanical homogenization of the tissue with a solvent

or mixture of solvents,
11) cold HCI acid dissolution of the biological tissue
followed by a partitioning into an organic solvent,
ill) solvent elution from a column of biological tissue

mixed with anhydrous sodium sulphate,
iv) steam distillation of biological tissue (in an aqueous
solution) followed by condensation of the water
vapour and trace organlcs and flow through partitioning
of the organlcs into a non-water soluble solvent.

The first three techniques, because of the presence of
high levels of lipids and fats, require the separation of these
coextractives from the extract prior to separation or
fractionation of contaminants and anlysls. The two most commonly
employed techniques for lipid - contaminants separation are:
1) acetonltrile - petroleum ether partitioning usually

followed by Forisil column chromatography (This

technique is referred to as the Mills. Onley, Gaither

technique).

- 490 -

11) automated gel permeation chromatography on Bio-Beads
SX-3 elutlng with dichloromethane or dichloromethane;
cyclohexane solvent mixtures.

EXPERHgWTAl SECTION

Fortified "clean" fish and unfortified "real" fish
samples were analyzed using three analytical techniques. Model
compounds, representing five chemical classes were used for
fortification at concentration levels of roughly 100 ng/g and 20
jLig/g. The "clean" sample was a lake trout from Lake Opeongo in
Algonquin Park and the "real" sample was a lake trout from the
north shore of Lake Ontario. Table 1 shows the model compounds
and surrogate standards used in this study while Table 2 shows
the concentrations of these compounds at the two fortification

levels.

For the analysis of the high level spiked samples 5

grams of tissue were extracted and the extract taken to a final
volume of 1.0 mL. For the low level spike and the unspiked
"real" fish 5-10 grams of tissue were extracted and the final
extract volume adjusted to 200 juL.

The three techniques evaluated are briefly presented

below:

Acid Digestion

Fish tissues were digested overnight In cold HCl and
extracted with 25% dichloromethane in hexane (v/v). Entrained
acid was neutralized with sodium bicarbonate and the extract
reduced in volume by rotary evaporation. Volumes were adjusted

- 491 -

TABLE 1
MODELCOmiMDS MD SURROGATE STAM)AROS

" CoS flTgfi

Level Level "Real"
Class Contaminant Spike Spike Sample

Volatlles Trichlorobenzene X X

Hexachlorobutadlene X X

Hexachlorobenzene X X

Chlorophenols 2-Chlorophenol 3,4,5,6-04*

2,4 Olchorophenol X X

2,4,5 Trichlorophenol X X

Pentachloropheno! X X

Aromatic
Amines

01 phenyl Amine

Organo
chlorines A
PCB

Ml rex

p,p'-ODE

Oleldrln

Aroclor 1242,54.60

X
X
X
X

X
X
X

Polyaromatic
Hydrocarbons

Fluoranthene
Phenanthrene 0-10
Anthracene 0-10*

X
X

Internal Standards

- 492 -
TABLE 2

SPHCIMG LEVELS OF FORTIFIED COIffOUIIDS

FOR LOW AMD HIGH LEVa SPIKIMG OF 'gEAII' FISH HOMOGEMTE

Low Level High Level

MODEL Spikes Spikes
COMPOUNDS "9/9 H3il±_

Dichlorophenol

Trichlorophenol

Pentachlorophenol

Di phenyl Amine

Trlchlorobenzene

Hexachlorobutadlene

Hexachlorobenzene

Mi rex

p,p-DDE

Dieldrin

Fluoranthene
Phenanthrene dlO

147

102

119

177

110

124

- 493 -
to 10 mis 1:1 dichloromethane/cyclohexane. A 1 ml. portion was
removed for lipid determination and a seven mL. aliquot was taken
for GPC. The GPC eluate collected was rotary evaporated to
approximately 2 mL., transferred to a reacti-vlal with rinsings,
with final volumes achieved by blowing down with a gentle stream
of nitrogen.

Polytron Homogenizatlon

Contaminants were extracted from the tissue using
dichloromethane, sodium sulphate and a polytron tissue
homogenlzer. The dichloromethane extract was reduced In volume
by rotary evaporation and made up to 10 mL. with 1:1
cyclohexane/dichloromethane In a calibrated centrifuge tube. A 1
mL. portion of the centrlfuged extract was taken for lipid
analysis with a 7 mL. aliquot of the remainder of the extract
processed through gel permeation. Once again, the GPC eluate was
rotary evaporated to 2 mL and quantitatively transferred to a
react1-v1al where nitrogen gas was used to gently evaporate to
final extract volume.
Steam Distillation

Fish homogenates were weighed and transfered to 50 mL
round bottom flasks with rinses of organic-free water. The
volume of water was adjusted to 300 mL. Three mL organic- free
water and 10 mL hexane were charged Into the condenser portion of
the apparatus. Steam distillation of the samples continued for 3
hr from the time distillation began. At the end of this time
heating was stopped and the system allowed to cool. The water
and hexane was drained from the condenser into a centrifuge tube.

- 494 -
The organic layer was removed and reduced to the desired final
volumn. The steam distillation extracts weree not processed
through the GPC prior to GC/MS analysis.

The Hpld removal step used throughout this study was
by automated gel permeation chromatography (GPC) on Bio-beads SX-
3 elutlng with 1:1 d1chloromethane:cyclohexane. Recovery studies
were pereformed on the GPC at the same levels tested on the fish.

The GC/MS analyses were performed on a Finnlgan 4510
using a 30 M SE-54 capillary column directly interfaced to the
ion source. Electron Impact (70 eV) spectra were obtained over
the mass range of 90-550 A.M.U. scanning once per second. The
column temperature program was 70^0 for 2 min.. 70^0 to 280^0 at
10°C per min. with a hold time of 10 min..

RESULTS AMD DISCUSSION

Fortification Studies

GC/MS analysis of samples processed using the polytron
homogenization procedure produced good mass spectra for all
compounds for both the low level and high level spikes. The acid
dissolution performed equally well for all compounds except
diphenyl amine which would not be extractable from the acid
digestion solution. Steam distillation failed to recover the
higher chlorinated phenols, and resulted in very poor recoveries
of diphenyl amine mirex, p,p'-DDE, dieldrin and fluoranthene.
The recovery data for the three methods are reported in Table 3.

TABLE 3
KCOlfERIES OF FORTIFIED CQITOUIDS

FOR LOU MR) HIGH LEVa SPIKING OF 'aEAH*

FISH HQHOGEIUTE

LOW LEVEL

HIGH LEVEL

MODEL
COMPOUNDS

Acid
Digestion

Recovery
t

Poiytron

HoMogenlzatlon

Recovery

Steam
Distillation
Recovery
X

Acid
Digestion
Recovery
%

Poiytron
Hoaogenlzatlon
Recovery
%

Steam
Distillation
Recovery
%

Dichlorophenol

Trichlorophenol

Pentachlorophenol

79.0
78.3
105

66.1
63.0
100

21.0

87.2
78.9
81.7

98.1
64.3
40.0

24.1
— 1

01 phenyl Amine

36.5

47.0

- ■

83.8

14.0 ^

Trichlorobenzene
Hexachlorobutadlene
Hexachi orobenzene

90.7
64.7
42.4

74,4
68.6
39.0

49.0
40,0
55.2

76,6
70.1
43.0

83.8
83.3
68.5

61,4
62.6
42.3

Ml rex

P»P-ODE

Dieldrin

36.8
22.2

44.7

44.4

4.5

23.4

83.1
94,7
79.6

93,0
86.8

3.5
9.9

12,3

Fluoranthene

Phenanthrene dlO

46.2

47.8

■ 43.8
43.3

19,3

60.0

93,9

88,4

92.6

81.4

6.3

40,3

High level in fish homogenate blank

- 496 -

Polychlorlnated biphenyls (PCB) were also tested for
recovery by the three methods. This data, reported 1n Table 4, shows
that while both the acid digestion and polytron homogenlzatlon
resulted In good recoveries for the dichloro- to heptachloro-
blphenyl, isomers the steam distillation recoveries decreased
significantly with Increasing chlorlnatlon and the overall
recovery of the PCB was less than 50%.

The results for the duplicate GPC spike recovery test
are reported in Table 5. In all cases, the recoveries from the
GPC were lower than the recoveries found In the overall fish
fortification, extraction and GPC studies- This has been
observed previously and seems to be related to the presence of
lipid materials acting as a "keeper" In all extract concentration

steps.

Figure 1 and 2 show the reconstructed Ion chromatograms

of the high level spike and low level spike respectively.
From the RIC plots, the peaks of the model compounds
are easily observed for the high level spike while the compounds
are far less apparent In the low level spike samples.
Foreground-background subtraction routines assist In further
defining these lower concentration compounds and in providing
useful mass spectra.

Naturally Contaminated Fish Study

The real fish homogenates demonstrated little
variability In terms of contaminants extracted by polytron
homogenizatlon or acid digestion. As in the fortification study,
the steam distillation extraction procedure was far less efficient In

- 497 -

TABLE 4
PCS RECOVERY

PCB TT? XcT^ Polytpon J^eii

Isomer Isomers Digestion Homogenlzatlon Distillation

^_____ % Rec % Rec X Rec

C12

74,3

65.5

91.8

C13

85.4

70.5

73.1

C14

93.5

85,7

35.9

C15

94.0

92.7

19.3

Cl6

102

103

4.7

CI7

103

Mean

92.0

86.7

37.5

Total PCB spike of 60 ^g

- 498 -

TABLE 5
GPC SPIKE KECOVERIES FOR MODEL COMPOUNDS

TSPc iJPT

MODEL COMPOUND Spike #1 Spike #2 Average

% % I

Dichlorophenol

Trichlorophenol

Pentachlorophenol

78.5
60.0
60.7

67.0
43.0
51.5

72.8
51.5
56.1

D1 phenyl Amine

86.0

74.8

80.4

Trichlorobenzene

Hexachlorobutadlene

Hexachlorobenzene

76.0
68.8
74.5

68.6
76.6

70.3

72.4
72.4
72.4

Mi rex

p.p'-ODE

Dieldrin

74.3
95.0
72.7

69.5
72.5
79.3

71.9
83.4
76.0

Fluoranthene
Phenanthrene dlO

93.0
82.8

84.8
71.3

86.9
77.1

FIGURE - 1 RIC HIGH LEVHL SPIKE

DATA: FPH2 tl
CALI: CALI2ee4 12

108. e-i

RIC

65-^/83 8:e2i9e

SAMPLE: 1 UL POLYTRON FPH2

CONOS.: SE54/D0/QEM

RANGE: G 1,1732 LABEL: N G, 4.9 QUAN: A 0, 1.0 J u BASE: U 20. 3

6 4

SCANS 300 TO 1732

RIC

576

751

1285

971

1042

802

T 1 r

1090

I 1202

1577

1475

670720.

T r

\^1 A 1691

FIGURE - :

RIC LOW LEVFL SPIKE

RIC DATA; P0LY2 #1261

85/20/83 11:08:00 CALI: CW.I0505 12

SAMPLE: 1 UL LOW LEl'EL SPIKE POLVTPON HOMOGENIZATION

CONDS.: SE54/IS'QEM

RANGE: G 1.1600 LABEL; H d, A,Q OUAN: A 0. 1.0 J t^JiE: U .

SCAB'S -00 TO 1600

100.0-1

1314

1229

564224.

RIC

248

^Vi,

■*^.

528 759 ,

lei?

o
o

- 501 -

It's overall performance. GC/MS results from the acid digestion,
steam distillation and polytron homogenlzatlon are listed In
Table 6. Specific Ion searches were also run for PAH and their
alkylated derivatives, chlorinated furans and dioxins, chlordanes
and each of the model compounds that were used In the
fortification study but no traces of these compounds were found.

The low recovery of the labelled chlorophenol
surrogate standard could be due to the low final extract volume
(200 ;j1) or as a result of deuterium exchange as has recently
been reported. Individual recoveries of this compound ranged
from 2.7 to 76%. Further studies are necessary to Identify the
causes of this variability. The d^Q anthracene gave consistent
recoveries averaging 106% for all methodologies.

The GC/MS outputs of the polytron homogenlzatlon and
acid digestion extracts contained rather large peaks that are
mainly fatty acids and hydrocarbons. Each successive GC/MS run
gave an increase in the total amount of material elutlng from the
glass capillary column. This Increase Is a carry over from the
previous GC/MS run. Much of the Interfering material can be
removed by base saponification, however It could also affect
compounds such as DDT. The steam distillation extracts produced
relatively clean chromatograms but quantltated only 25% of the
contaminants Identified compared to the other extraction
procedures.

PCB were observed In the extracts produced by all three
methods. The concentration of total PCB was 0.60 + 0.07 jug/g for
the steam distillation 0,80 to .05 uglg for the polytron

TABLE 6

cqrTJwiMWffs iPomFicp » km. fish hqnkemtis by bc/ns

("B/BB)

Mean

Cont«iilMnt Concentration

Add
Digestion

Relative Recovi
Polytron
Hoiiogenlzatlon

try

CONTAMINANT

Add
Digestion

Polytron
HoMogenlzatlon

Steam
Distillation

Steaia
Distillation

Hexachl orobenzene

25.6

21.2

28.5

100

Pentachlorophenol

67.9

100

DOE

290

209

78.5

100

Octachlorostyrene

7.5

7.2

100

t-nonachlor

9.7

10.5

100

- o
to

DOT

143

97.7

13.5

100

Ml rex

37.0

33.9

10.2

100

Photoni rex

8,7

8.8

2.5

100

Recovery (%)

D-10 Anthracene

111

110

90.1

100

D-4 Chlorophenol

5.4

100

- 503 -

homogenlzation and 1.05 + 0.06 ;jg/g for the add digestion
procedure.

The data obtained from the fortification studies and
the analysis of naturally contaminated fish tissues show that the
steam distillation tehcnique Is clearly Inferior In the
extraction and recovery of high boiling toxic organlcs from fish
tissue samples. The acid digestion technique produced slightly
higher recoveries of the chlorophenols, polycycllc aromatic
hydrocarbons, volatile chlorinated hydrocarbons, and PCB in the
fortification studies.

In addition, the acid digestion technique resulted in
the higher contaminant concentrations for 75% of the compounds
quantltated in the naturally contaminated tissues and was used
for all subsequent tissue preparations.
HDMITORIIIG KESULTS

The species and location of fish samples monitored are
sunmarlzed In Table 7. All samples were analysed by GC/MS for
the purpose of Identifying organic contaminants, with special
emphasis In Identifying compounds that are not presently
monitored by MOE. Table 8 lists the compounds that MOE routinely
monitors. Tables 9 and 10 report the compounds that were
Identified by EI and NCI GC/MS respectively. These latter tables
offer a comparison between EI and NCI analyses. These results
also demonstrate the capability of NCI to detect compounds such
as heptachlorostyrene, toxaphene, PCDPE's and pentachloroanlsole,
compounds that gave minimal or no response on their EI
counterparts.

- 504 -

TAHLe? . SPECIES AND LOCATION

SPECIES

CHANNEL CATFISH
YELLOW PERCH
AMERICAN EEL
WHITEFISH
L.AKE TROUT

TOTALS

SUPERIOR

HURON

ERIE

ONTARIO

NONE

2 OPEN

3 OPEN

NONE

1 OPEN

2 OPEN

3 OPEN

2 OPEN

NONE

1 OPEN

NONE

2 OPEN

NONE

HOT SPOT

NONE

3 OPEN ?<
i HOT SPO

1 OPEN !?<

6 OPEN

6 OPEN ^.<

HOT SPOT

1 HOT SPO

TOTALS

5 OPEN

8 OPEN

1 OPEN

2 OF-EN

3 OPEN ?■<
4 HOT SPO"

19 OPEN S<

- 505 -

TABLE 8. MOE COMPOUND SEARCH LIST
FOR SPORT FISH

POLYCHLORINATED BIPHENYLS

OCTACHLOROSTYRENE

p,p '-DDE

p,p '-DDD

p,p '-DDT

o,p -DDT

HEXACHLQROBENZENE

a-BHC

b-BHC

g-BHC

ALDRIN

HEPTACHLDR

a-CHLORDANE

g-CHLORDANE

MI REX

TOXAPHENE

- 506 -

TABLE 9

COMPOUNDS IDENTIFIED BY ELECTRON IMPACT (EI)

COMPOUND

THUNDER BAY SAULT~STE-MARIE
LAKE TROUT LAKE TROUT

BURLINGTON PORT WELLER
LAKE TROUT LAKE TROUT

DICHLOROBENZENE

TRICHLDROHENZENE

TERACHLOROBENZENE

PENTACHLORQBENZENE

HEXACHLOROBENZENE

p,p '-DDE

o,p ' -DDE

p ,p - DDT

p ,p '-DDD

o,p '"DDT

DDMU

BENZOIC ACID

MI REX

PHOTDMIREX

CHLORDANE

NONACHLOR

CL3~TERPHENyL

CL4-7ERPHENYL

PCB CL2~CL8

DIE^ROMGPHENDL

C 2 -PHENOL

C3- PHENOL

C4-PHEN0L

C8- PHENOL

NONVL PHENOL

NAPHTHALENE

METHYL NAPHTHALENE

PHENANTHRPNE

BHC

DIELDRIN

TETRACHLOROCYCLOHEXANOL

CHl-.OROi:,{UTYLTIN

T R I METHYL BENZ ALDEHYDE

CHI ORGBROMG TOLUENE

HEPTACHLOROSTYRENE

OCTACHl GROSTYRENE

C2-THI0raLUF.NE

C4-TH 10 TOLUENE

UNf.

1 90

CL2

UNK

2^6

CL4

UNK

2 1 6

UNK

170

UNK

116

CL 1

UNK

tt6

242

CLl

UNK

tt7

174

CL2-3

unk:

ttB

192

CL2

ND
ND

N0
HD
ND
ND
ND

nd:

ND
ND
N0
X

m
m
m

X
X

ND
NO
ND
ND
N0

- 507 -

TABLE 10
COMPOUNDS IDENTIFIED BY NEGATIVE CHEMICAL IONIZATION (NCI)

COMPOUND

CL3-NI TROBENZENE

CL4-NITR0EIENZENE

HEPTACHLOROSTYRENE

OCTACHLOROSTYRENE

CHLORDANE

NONACHLOR CIS ?< TRANS

TOXAPHENE

MIREX

PHOTOMIREX

PCDPE CL6-CL7

PCB CL5-CL9

DIBROMOPHENOL

TRIBROMOPHENOL

DIELDRIN

PENTACHLOROBENZENE

HEXACHLOROBENZENE

PENTACHLDROANISOLE

C14 HI 8 O CL4

THUNDER BAY SAULT~STE-MARIE BURLINGTON PORT WELLER

LAKE TROUT LAKE TROUT

X(2)

X(2

- 508 -

A suninary of the compounds that were detected and are
not being routinely monitored is given in Table 11 along with the
fish species and the lake of origin. Mirex Is Included in the
list to emphasize it's detection by GC/MS analysis in Georgian
Bay. It's concentration was calculated at 4.7 ng/gm. Few of the
compounds listed were above 50 ng/gm and those compounds that
exceeded that threshold are presented In Table 12. Of these
compounds p,p'-DDE and PCB's ranged from 50 ng/gm to 2 >ig/gm
while compounds such as the chlorobenzenes ranged from not
detected to 5 ng/gn.

Figures 3 to 7 show the mass spectra of selected
compounds that have been Identified in the current study. These
include, by NCI; tribromophenol from a Sault Ste. Marie lake
trout (Figure 3), heptachloro diphenyl ether from a Burlington
lake trout (Figure 4), and Heptachlorostyrene from a Port Weller
lake trout (Figure 5), and by EI; a butylated tin from Jordan
Harbour lake trout (Figures 6 and 7). Figures 8 and 9 depict the
reconstructed gas chromatograms (RGC's) from EI and MCI for a
Port Weller lake trout. The sample extract was treated to base
saponification prior to analysis which removed much of the fatty
acid compounds that are prevalent in fish extracts. The two
RGC's reveal the different outputs that emerge from EI and NCI
techniques. Octachlorostyrene, scan 1146, from the NCI RGC Is a
major contltuent of the chromatogram while under EI, It is barely
discernible, being almost masked out by a co-eluting hydrocarbon
as seen in the EI spectra shown in Figure 10. In comparison.

- 509 -

TABLE ll.NEW COMPOUNDS IDENTIFIED BY EI/NCI-GCMS

COMPOUND

DICHLOROBENZENE

TRICHLOROBENZENE

TETRACHLOROBENZENE

PENTACHLOROBENZENE

PENTACHLOROAN I SOLE

HEPTACHLORQSTYRENE

PHOTOMIREX

DIELDRIN

NONACHLOR

o,p'-DDE
DDMU

C2-PHEN0L

C3-PHEN0L

C4-PHEN0L

C8-PHEN0L

C9-PHEN0L

TETRACYCLOHEXANOL

DIBROMOPHENOL

TRIBROMOPHENOL

CHLOROBROMOTOLUENE
CI -3 TERPHENYL
CI -4 TERPHENYL
CI -6 DPE
CI"? DPE

Cl-3 NITROBENZENE
CI -4 NITROBENZENE
C14 H8 C14

BENZOIC ACID

NAPHTHALENE

METHYL NAPHTHALENE

PHENANTHRENE

C2-THI0T0LUENE

C4-THI0T0LUENE

CHLOROBUTYLTIN
MIREX

SPECIES

LAKE

WF , CC , YP

S,H,E

S,0

YP,LT

S,0

YP,LT

YP,CC

H,E

LT,CC,YP,WF

S , H , E ,

WF,GB

H,0

LT,CC,YP

H,E,0

S,0

5,0

S,0

S,H,0

S,0

CC,YP,WF

H,E

LT,WF,CC,YP

S,H,E,0

LT,CC.YP

S,H,E,0

- 510 -

TABLE 12. COMPOUNDS IDENTIFIED AT 50 PFB OR GREATER

COMPOUND

DICHLOROBENZENE
TR I CHLORODENZENE
TERACHLOROBENZENE
HEXACHLOROBENZENE
p,p'-DDE
p,p '-DDT
p,p '-DDD
o,p -DDT
BENZOIC ACID
MI REX
CHLDRDANE
NONACHLOR
PCB CL.2-CL8
C2-PHEN0L
C4-PHEN0L
BHC
CHLOROBUTYLTIN
TRIMETHYL BENZALDEHYDE
OCTACHLOROSTYRENE
C2-THI0TaLUENE
C4-THI0T0l_UENE

FIGURE 3

TRIBROMOPHENOL - NCI

iee.e-1

se.e-

HID HASS SPECTRUM
12/1V83 14i34i« * 6:26
SAMPLE: SAULT STE MARIE MCI
CGNDS.i DeVOO/OEH
GC TEMP; 189 DEC, C

DATA: SSnSlHCI^ 1313
CfiLU CM.ieil2 «3

BASE N/Et 332

RICi 764328.

33 .7

176. e

148.1

lu.e

M/E

161

166

127.6

< , I 127

219.2

235.6

262.1

-?9¥4.

256

386

I •■il""

r 147436.

FIGURE *

HEPTACHLORO DIPHENYL ETHER - NCI

10e.0n

HID hASS SPECTRUM

08-^31/33 17il4!e0 + 24:57

SftflFLE: 1 UL POR.T WELLER LAKE TROUT

cores.: DBl/00/QEM

GC TEMP: 261 DEC. C

ENHANCED (S 15B 2N 2Tj

DATA; PMLTNCi:: 11361
CALI: CALIlCCe 13

BASE n/£: 376
RICi 10915e0.

376

r 203488.

f^A Q-

se.e

195

34a

366

2'jf

rV'E

^t"?

T 1— r

■ m ! ■

234

' T T T '

:>.fl

;:uy

320

J2D

'tTI'TIIIIFFIII

■40

365

■>jy

RGURE 5

HEPTACHLOROSTYRENE NCI

MID MASS SPECTRUM

08/31/83 17114:90 + 19:23

SAMPLE: 1 UL PORT HELLER LAKE TROUT

CONDS.: Oei/DO/QEM

GC TEMP: 217 DEC. C

OATAi PULTNCI^ 11857
CALIt CALIIGW 13

BASE M/Ei 274
RICi 239040.

100.0-1

273.8

50.0-

239.8

180.1

214.1

M/E

180

194,0

I""''

♦4

248.1

264.2

200

220

240

■ • ■ I ■
2S0

r 83872.

T»-

389.8

280

300

320

343.7

340

FIGURE 6

SCAN #676

JORDAN HARBOUR. LAKE ONTARIO

\m.%n

58.0-

mss SPtcmm mtai jhyip igtg

01/t9/84 16i39>08 * 111 16 CALIt CALieteS tl

StfTLEi lUL JORDPN HAftBOUR YELLOU PERCH COTPOSITE FU-2eeU.
CONDS.i dB^lS/i3B\
GC TE»>: 179 DEC. C

W/E

213.8

231.8

■ ' I •■

BASE H/Ei 2E9
RICt 17929.

269.1

r 1214.

t— 1
I

291.3

FIGURE 7 LIBRARY MATCH SCAN #676 - JORDAN HARBOUR

LJ^fY «WJ MTAi JHiriP • g7S

9X^m/9i IClMiM ♦ lit 16 CM.It CM.Itli9 • 1

SMPUi lUL JORMH HMMUR YELLW PERCH COrOSITE RMMUL
C0ND5.I On/IS/«H
DtMCED (S 19 » ST)

BASE n/tt 269

RICt 9313.

U«

C12.K27.a!sN '

C12.H27.CUSN

, ■ , ■ , ■ ,1 .all I , I , w y

STMMVC. CIUMrmiS(2-ICT»fVLPR

1-4-1 . ^ , , , — UU^

STMMNIE, TRIBOTYLOLORO-

4-THIAZ0LBCETICMCID.2-<IMXi)

I I . I ■ m I u I I I J

U-, ...i l ls ^--^ , *JWi

.-jjii

, II

>i ■ . " ^i

I ■ »

' — ■ ■ 'i' ■ ■' — i — ■ -^i ■ I I- » ■ iii. ■ |l , , , , ii, , , ,I| . , . ,

in 19t 2n 29B

FIGURES

RGC PORT WELLER LAKE TROUT - NCI

iee.0-1

MID RIC OATAi Ptt-TNCi;! «1 SCANS 588 TO 1880

08/31/83 17jl4:ee CALI: CALie929 t3

SfWLE: 1 UL PORT WELLER LAKE TROUT

COHDS.: Oei/DO/QEM

RfWGE: G 1.2698 UCEL: H 8, 4.8 QUAH: A 8, 1.0 J 8 BftSE: U 20> 3

RIC

592

L-L

&e0

11:08

392

729

846

781

888
14:48

1886

964

1309

1228

U46

1275

1470
I

1808
18:28

7495678.

1288
22:88

1688
29:28

1388 SCAN
33:08 TII€

FIGURE 9

RGC PORT WELLER LAKE TROUT - £1

lOe.0n

RIC DATA: PWLTSAPON #1

10/87/33 14:04:00 CALI; CALI1007 «3

SAMPLE: lUL FW-LT BASE SAPOHIFIChTIOH F.U,= 400 UL

CONDS. : Dei/DO/GtEH

RhNQE: G 1.1691 LABEL: H 8. 4.0 QUAN: A 0. 1.0 J BASE: U 20. 3

SCANS 275 TO 1631

RIC

384

317
.1, Lui

552

^^^^^ f. ■,..

6:40

10:80

909

800

13:20

1000

18:40

1208
20:00

1400
23:20

1581

140800.

1637

•si

1688
26:40

scm

FIGURE 10

OCTACHLOROSTYRENE - £1

mss sPEcmjn

•9/31/83 ni47in * 18i«l

SMVLEi I UL PWrr tCLLER UlU TROUT F.U.-GW UL

CQMS.i DB1/X/Q01

rc TDTi 237 DEC. C

DMMCED <S 19 2N ST)

tMTAt fMLTK IIMI
CM.Ii CM.I2211 13

BASE M/^i 97
RICi 1648e.

loe.e-i

se.e-

lee.e-t

58.8-

n/E

181

188

256

1 4

l?3

120

137

151

166

148

160

180

r 18M,

ii,;aiif^fi..,iiiiii.',^,'.iii,/.i?..,, X' ,^i'k T ^-

I— >

248

r 1894.

268 , 282

^ 'i I I I ■ ■ ■ M | ■'■ I ■ I I I I ■ I I ■ ■ I I I

386

3*3 353

388

268

288

388

328

■TTf'^ I'f'l 1 I 1*1 I I I I I I I I I I >■

346 368 388

.v..

- 519 -

octachlorostyrene under NCI conditions shown In Figure 11, Is
will defined with little interference.

Figures 12 to 16 are ion plots of toxaphene, chlordane
and nonachlor for a Thunder Ba^y lake trout extract and a standard
under NCI conditions. All these compounds are easily Identified
while toxaphene, under EI conditions, produces no response
whatsoever and chlordane and nonachlor give relatively weak
responses. Figures 17 to 20 depict the spatial distribution of
selected organic contaminants for each lake. Few levels are
above 50 ng/gm. Significant levels of the chlordanes and
nonachlor are seen In a whitefish and a channel catfish In Lake
Huron, averaging 85 and 120 ng/gm respectively. Two channel
catfish from Lake Erie give values of chlordanes and nonachlor
averaging 82 to 120 ng/gm respectively. In Lake Ontario, ml rex
Is reported as high as 95 ng/gm In one lake trout. As mentioned
previously, other compounds produced signlflcont levels but were
not plotted. Figure 21 compares contaminant levels, In lake
trout samples from Lake Superior versus lake trout samples in
Lake Ontario. Only p,p'-ODE was determined to be consistently
above 50 ng/gm concentration. Interestingly, BHC levels in Lake
Superior are equivalent to levels In Lake Ontario, while p,p'-DDE
is significantly higher in Lake Ontario, averging 350 ng/gn for 4
samples while. In Lake Superior samples, a mean of 80 ng/gm for
samples was observed. Mirex and phoromlrex were noted in all
Ontario lake trout while they were not detected In any of
Superior's lake trout samples.

FIGURE 11

(XTTACHLOROSTYRENE - NCI

niD MASS SPECTRUM

08/31/83 17;14:M + 21:91

SAMPLE: 1 UL PORT WELLER LAKE TROUT

C0ND5.: DBl/DD/QEH

GC TEMP: 230 DEC. C

i00.e-i

MTAi PHLTHCI7. #1146

CALii CALueee #3

387,8

BASE M/E: 308
RICl 3575800.

50.0-

273.9

150. 1

212.^

' I ' ■

i-i-b-.y-

i;J, 24i.y

i I ! ;ii ,1

tlliU

289.9

r 794624.

**7

379.

fl'E

2CiO

J.tiM

300

— I — • — : — ' — r

350

FIGURE 12 ION PLOT MASS 3»3 A 377 TOXAPHENE STANDARD - NCI

MID Ric + mhss chrom^togrhms oath: tokchlor:: n

12/15^-'33 lC:02:i3i3 TALI; rALI2211 **:*-:

SAMPLE: TOXhPHEHE AMD CHlCt'^AK^ '"ThKDmFD
COMDS.: DB5.-DLi/uEf1

RANGE: G I, S4? LhEEL: N 0.- -^.O QIJhM: h U.- 1.0 J ShSE: U 20. 3
84.8-1 ^-^^

SCANS 600 TO 947

130.3-1

37?

RIC

S5:i

713

;^5

7S5

776

826

. 7q?: ft S43

318

T — *r

< i
I ;

:-:i:< PT-=; c-40 /"'j':

bib

31712,

342.768
'< ± 8a566

37376.

376.8%

± 0.5ee

2404358,

500
7:31

T ■■ T

k;f

1-;

930
1!:1E

TIME

FIGURE 13

ION PLOT MASS 3*3 4 377 THUNDER BAY LAKE TROUT - NCI

33. 3n

MID RIC ♦ MASS CHROhATOGRAMS DATA: TBLTHCi;! 1765
12/15/83 3:30:06 CALI: CALI22n 13

SAMPLE: THUNDER BAY NCI FM=2ee UL
CONDS.: DBS/DO/QEM

RftNGE: G 1. 966 LABEL: M 6. 4.0 GUAM: A 0, 1.0 J Q Eh_-E: U 20.-

765

SCANS 696 TO 966

lee.en

361E0.

342.700

± 0.500

376. 590

± 0.500

34O7S70.

686
7:31

1 — I — I — r

see

10:01

850
10:39

900
11:16

95S
11:54

SCAH
TIIC

FIGURE 14 ION PLOT - MASS »10 A »»» CHLORDANE STD - NCI

109,0-1

410

MID RIC + riHSS CHROMhTOGRAMS DhTh: imiHio^z #1

12/15/S3 10:02:00 CiCiLI; CALI2211 #^

ShMPLE: TO:^"hPHEilE hND CHLGi=LiHfiE SThMDmFD
CuNDS.: DB5.DD/C!Eri

PAUGE: G !.■ 347 LhBEL: N 0. 4.6 OUihN: h 0.. 1.0 J BASE; U 'Su

715

727

£C«13 bOO TO 347

■5.5-1

444

794. G-

RIC

bo7

. 1^, 1—

-r-'-^

::i

' i '

7S1

t^y

f.'-^-

Zli.

■-■ ■ —

—■■ _•- -.-.

- --■ .■„• ■-- ■-- -

' ■

■

7l\

I !

■ ■ r j
' ' '

S02

T 1 1 r

S54
— — r

926

T r^;-i r

T r

T 1 1 r

?o7

?1G S2o ^-^J"-,

904

603
7:31

1 H.i7-i

1 1 r r

900

11:16

302592.

409.766

1 o.see

228£e8. ^

443.886
1 6.566

^64356.

SCAN
TIMt

nCURE 15

ION PLOT MASS 410 THUNDER BAY LAKE TROUT - NCI

100.0-1

MID RIC + nftSS CHROMftTOGRAM DATA: TBLTMCI7. #732

12/15/S3 9::30:ei3 CALI: CALI22n #3

ShMPLE: THiJtCiER BAY NCI FU=2yti UL
CmMDS.: DE'5/OD/QEM

Pp:^::E: G 1. ^56 LhB^L: H 0.- 4.9 QUAH: h d.. 1.^3 J Q BhSE: U 20.

723

scw^ Eoo TO see

4ie

RIC

6GS

T" T I

688

eG3

63t>

73G

^^745 764^ jl^ . ^ 7,-1^ ^y, ^A^

331

902 ^F ^ ^.^

951

--"V^-'-'

ai'vj'"W'^^^''

■877

887

*'"UJ)JI

43403.

409.760
± 0.583

3407370.

T 1 r

T r

T 1 1 r

500
7:31

8:89

700
8:46

750
9:24

I ■ I — 1 r

800

16:01

T "-r — 1 1 ■ 1

350
10:39

T 1 r

900
11:16

11:54

SCAH

THE

m^am - vm^m-- .t -itv^Jt. mrt^'.-f ■,*m.:tm-if*ritj .:^».

''i':r\- '--T-

FIGURE 16 ION PLOT MASS »»» THUNDER BAY LAKE TROUT - NCI

100.0-,

MID RIC + MhSS CHROMATOGRhM DATA: TBLTNCI^ 1732

12/15/83 9:30:00 CALI: CALI2211 «3

SAMPLE: THUHDER BAY MCI FU=2e0 UL

CuHDS.: DBS/DO/QEH

FhM'JE: U i. 965 LAEEL: N U, 4.e__QIJAN: A Q. 1.0 J

781

444

RIC

603

T — r

b63

t.Hb

bS5

SCAMS E0O TO 966

BhSE: U 20/ 3

732

-' •< '^^\

WV Wv]^ 'a

961

^ — "-

737 SOS

3 7

79b

353

y3j

i^-! ,»wmH

877
I

887

153600.

443.800
± 0.500

T r

949

3407S70,

935

v.^.^j ul/', r..-.., 919

1 — I 1 — r

600
7:31

~1 — ^

650

8:09

1 — I — r

700
8:46

-I 1 r

750

9:24

~-\ — '

S00

ie:ei

1 1 1 r

350
10:39

900

11:16

T — '
950
11:54

SCAM
TIME

FIGURE 17

SPATIAL DISTRIBUTION OF SELECTED CONTAMINANTS IN LAKE SUPERIOR

2
O
\

H
Z
UJ

O
O

LAKE SUPERIOR

MOE SPORT FISH 1983-84

kti lJ MM MM

IZZl HX

LAKE

BH ^ DE

TR OUT- LOCATION

^ DD KZl CH

^^^
\;^^

^ NO

FIGURE IS

SPATIAL DISTRmUTlON OF SELECTED CONTAMINANTS IN LAKE HURON

z
o

bi
U

z
o
o

150

140 -
130 -
120 -
110 -
100 -
90 -

ao -

70 -
60 -
50 -
40 -
30 -
20 -
10

im^

WF-GB

ZZl HXB

WF-S

rr^ BHc

LAKE HURON

MOE SPORT FISH 1983-84

CC-S

YP-GB

YP-S

LOCA TION ^SPE CIES

^ CHD ^S NGN K3 MIR

FIGURE 19

SPATIAL DISTRIBUTION OF SELECTED CONTAMINANTS IN LAKE ONTARIO

2
O

bl
O

o
o

cc-stc

LAKE ERIE

MOE SPORT FISH 1983-84

CC-W

CC-C

YP-W

[771 HXB

SPECIES ^LO CATION

KS BHC ^ CHD

YP-C

^ NON

YP-C

nCURE 20

SPATIAL DISTRIBUTION OF SELECTED CONTAMINANTS IN LAKE ERIE

2
O

hi
U

z
o
o

100

LAKE ONTARIO

MOE SPORT FISH 1983-84

[ZZI HX ES BH

LAKE TR OUT- LOCATION

UP7X CH ES NO KZl MX ^S PX

FIGURE 21

- S30 -
SPATIAL DISTRIBUTION OF Sr.L'-CTION CONTAMINANTS LAKF: SUPERIOR vs LAKE ONTARK

•O

50-

40-

90-

20-

10-

LAKE TROUT
Moc spom- nsH ims-«4

s
s
s
s
s
s
s

s
s
s
s
s
s

s
s

s
s
s
s

ZZI HXi

■HC

ESa CHD ^SImon

460

400-

950 -

900-

290-

200-

150-

100 -

50-

LAKE TROUT

MCE SPORT nSH 19S9-B4

LAKE 80PEI

CTTl p,p*-DDE

MIKEX

l/KE

ONTARIO
PHOTMtRCX

FIGURE 22

SPATIAL DISTRIBUTION OF SELECTED CONTAMINANTS FOR YELLOW PERCH IN GREAT LAKES

240

YELLOW PERCH

MOE SPORT FISH 1983-84

RIOR

HURON

ERIE

Ontario

ZZl P.P'-DDE

LOCATION

KS CHD ^ NON

^S3 HXB

FIGURE 23 SPATIAL DISTRIBUTION OF SELECTED CONTAMINANTS FOR CHANNEL CATFISH IN LAKES HURON AND ERIE

O
\

I-
<

U
O

o
o

150

CHANNEL CATFISH

MOE SPORT FISH 1983-84

s stc

LA KE HURON

[771 ODD r^N] CHD

LAKE ERIE

^2m NON

- 533 -

Figures 22 and 23 are similar outputs for yellow perch
and channel cat fish respectively.

Noteably the highest levels, for those compounds
plotted, were seen 1n the Lake Superior yellow perch.

SUWWRY

Few of the contaminants that were Identified exceeded
50 ng/gm. PCB's and p,p'-DDE were the most predominant
contaminants, being present 1n all fish samples. Although many
contaminants were observed at below 50 ng/gro. It does not Imply
that their reported levels are not significant, and ndy Indeed be
cause for further investigation.

NCI offered a good means for detecting higher
halogenated organlcs that were not detected by EI but was limited
to this class of contaminants. A limited number of fish samples
have been extracted to date, and these represent approximately
jO% of the total number of fish that will be processed under this
current study. The completion of the remaining samples will
hopefully add considerable information to the overal 1 view of
sport f1sh in the Great Lakes.

- 534 -

Canada-Ontario AjrecMnt Report (1981). tnvlronnental Baseline
Report of the Niagara River: Novenber 1961 Update*.

Cobum, J.A. and CoKba. M.E. (1981). "Identification of
Polychlorlnated Dipenyl Ethers In Whitby Harbour Bottoia
Sediments*. Presented at the A.O.A.C. Spring Workshop,
Ottawa, Ontario, May 1961.

Elder, V,A., Proctor, B.L. and Hites, R.A. (1981). "Organic

Conpounds Found Near Ouiap Sites In Niagara Falls, New York",
Environ, Scl. Techno!. 15^, (10), 1237.

Harless, R.L. and Lewis, R.G. (1980). "Quantitative Capillary
Colunn Gas Chro«atography - Mass Spectronetry Methods of
Analysis for Toxic Conpounds", Presented at the Pittsburgh
Conference on Analytical Chemistry, Atlantic City, New
Jersey March, 1960.

Kesselberg, RJ. and Seelye J.G. (1982). "Identification of
Organic Compounds in Great Lakes Fishes by Gas
Chro«atography/Mass Spectrometry", U.S. Fish and Wild Life
Service Administrative Report No. 82-1.

Kuehl, O.U. (1961). "Unusual Polyhalogenated Chemical Residues
Identified in Fish Tissue fro« the Environment", Cheaosphere
10, (3), 231.

Kuehl, D.W., Dougherty. R.C. (1979). "Screening of Huaan and
Food Chains Sanples for Contamination with Toxic Substances
using Netaglve Chemical Ionization Mass Spectrometry", In
press Advances In Mass Spectrometry , Volume 7.

- 535 -
KEFEWiCES CIWTIWEP

Kuehl, D.M. and 0ou9herty, RX. (1980). "PwUchlorophtiwl In
the Environnent: Evidence for Its Origin from Coaverclal
Pentachloroptienol tMi Megctove Chealcal Ionization Mass
Spectroawtry", In press Environ. Scl. and Techno!.

Ryan, J.T. Lau, P.Y., Pllon, J.C, Lewis, D., McLeod, H, Calway,
p. and Gervals, A. (1982). "^.a.T.e-Tetrachlorodlbenzo-p-
dloxln (TODD) Incidence and Levels In Lake Ontario
Coiaaierlcal Fish", presented at the 184^^ National A.C.S.
Meeting In Kansas City, Missouri, September, 1982.

Yurawecz, M.P. (1979). "Gas Chroaiatographic and Mass

Spectroaetrlc Identification of Chlorinated Trifluorotoluene
Residues In Niagara River Fish", J. Assoc. Off. Anal. Chem.
62, (1), 36.

- 537 -

HEAVY METAL MOBILIZATION AND BIOLOGICAL UPTAKE;
COBALT NINE TAILINGS

E. Hanna

JE Hanna Associates Inc

Abstract

The objective of the study was twofold. The first was to
assess the potential and significance of heavy metal leaching
and uptake from mine tailings in Cobalt, Ontario and similar
biophysical environments. The second was to explore the
impact that construction (on the tailings) of an artificial
marsh designed to treat municipal sewage might have on heavy
metal cycling .

The primary element of the study was a series of simulated
leaching experiments using cores of tailings to observe the
effects of i) ionic strength of the leaching solution, ii)
water table elevation and iii) redox levels on metal
mobil ization. The results of the experiments were
corroborated with field data. Biological uptake and cycling
was assessed using cultures of duckweed (Lemna minor ) .

The results of the work demonstrated that a large range in
the mobilization rate of heavy metals occurred over the
treatments used for the experiments. in addition, most
metals were in highly available forms.

- 538 -

INTRODOCTION

This project addresses both a known problem and one which
could arise in the near future. The known pr obi era is that of
contaminant leaching from mine tailings in particular heavy
metals (Down and Stocks, 1978) . Associated with this
leaching is the pollution of both ground and surface waters
and the effects on both the health of people drinking the
water and the environmental effects on other biological
organisms. Considerable effort has been directed to
controlling contaminant leaching from tailings but a detailed
understanding of chemical mechanisms of contaminant release
and biological uptake is lacking {Jenne and Luoma, 1978).

The tailings in the Cobalt area (Figure 1) are unusual in
that they are quite alkaline al though a large range of heavy
metals in relatively high concentrations are present. The
neutralization of acid tailings is a common initial step
pr eced ing vegetative reclamation (Peters, 1978). The Cobalt
tailings offer an opportunity to examine the behaviour of
heavy metals and biological uptake in a quite different
chemical environment from the more typical acid tail ing s
associated with sulphide ore bodies. Concomitantly, the
mitigation techniques appl i cable to alkaline tailings may
well differ substantial ly from those suitable for acidic
wastes ,

The second potential problem which this project addresses is
imminent and relates to an ongoing project funded by the
Ontario Ministry of Northern Affairs and coordinated by the
Ontario Ministry of the Environment , Two experimental
artificial marshes which receive municipal sewage effluent
have been constructed at Cobalt to test the feasibility and
performance of this technology; one marsh is built on
tailings. If the systems are effective, a full-scale marsh
will be built to treat sewage from Cobal t ,

The population is currently serviced by municipal water
supply and sewer systems but there is no sewage treatment
facility. The sewers outfall at two locations and directly
enter a stream system that flows out of the Town to the east
(Figure 2) .

The primary focus of the experimental marsh is on the
treatment of sewage; however , changes in the groundwater and
soil chemical regime, in addition to the introduction of
marsh biota on the tail ing s may significantly alter the
mobility and bioavailabil ity of contaminants (Allen and
Unger , 1980) .

/ Drainage

' Ditch ,

^ -^'^ TAILINGS

COBALT

FIGURE 2

- 541 -

The overall study objective was:

To determine the ■obility and bioavailability of heavy
metals in sine tailings under varying moisture and
chemical reg imes .

These results are of direct relevance to tailings management
in particular in the Cobalt area. Specifically, the study
was intended to derive estimates of:

i) Potential changes in leaching rates and biological

uptake of contaminants from tailings if a full-scale

artificial marsh system were constructed;
i i) Current leaching rates and biological uptake of

contaminants under varying edaphic cond i tions ;
iii) Changes in leaching rates and biological uptake of

contaminants that can be expected wi th disturbance to

the tail ings;
i V) Likely effects of mitigation strategies such as

revegetation, altering water tables, and chemical

treatments on leaching rates and biological uptake of

contaminants.

The research consisted of essential ly two components, i)
f ield sampl ing and ii) laboratory experiments . The emphasis
wa s on the latter component al though each served an essent ial
role. This paper deals only with the results of the
laboratory leaching experiments . A report describing all
aspects of the study has been submitted to the Ontario
Ministry of the Environment (JE Hanna Associates Inc., 1984).

The laboratory experiments centered on intact cores of
tail ings that had been collected from four types of sites at
Cobalt namely:

i) dry, sparsely vegetated , unenr iched;
ii) wet, vegetated, unenriched;
iii) wet, vegetated, enriched;
iv) recently d isturbed and flooded , vegetated ,
enr iched .

The columns were leached with various solutions to see how
metal release rates were affected . In anal yz ing the metals
released, bioavailability was determi ned us ing a bioassay
type of procedure.

- 542 -

METHODS

A ppn rritu ^ S ot Up

Tntrict rores of trail incjs were colloctoii durin<i the first
riolii survey in m itl-Oc tober . Plexiglass tubes, 6.4 cm in
diameter, were manually forced into the tailings and then
drawn out. The ends of the tubes were sealed with duct tape
and stored in a upright position.

In the laboratory, plexiglass plates with spigots were glued
to the base of each column and they were then mounted on a
retort rack with the upper surface of the tailings in each
column level with one another (Figure 3). Tygon tubing was
used to conduct the leaching solution to the top of the
columns and the leachate from the columns to sealed
Er lenmeyer flasks. Plates of plexiglass with holes drilled
to the diameter of the tubing were placed on top of the
columns to minim ize evaporation and contamination.

A common reservoir held the leaching solution which was
pumped to the columns with a peristaltic pump.

Leac hing Solutions and Conditions

A series of leaching solutions and conditions were used to
assess the mobil ity and bioavailabil ity of heavy metals in
the cores. They were selected to represent the types of
changes that are anticipated if an artificial marsh sewage
treatment system were to be built on the tailings.
Specifically, two solutions were used, i) neutral dilute
unenriched water and ii) neutral higher conductivity
enriched water. The dilute solution represents the natural
leaching process (eg, rain). The other solution is typical
of sewage effluent in Cobalt (G. Miller, pers.comm.). The
characteristics of the two solutions are shown in Table 1.

Three conditions were combined with the leaching solutions,
namely i) watertable below surface (ie, 20 cm), ii)
water table above sur face ( ie, 2 cm) and aerobic, and
iii) watertable above surface and anaerobic.

The anaerobic condition was simulated by using a layer of
vegetable oil on top of the surface water to prevent surface
oxygen exchange.

Combining the two leaching solutions and the three leaching
conditions resulted in six treatments in total.

- 543 -

SOLUTION RESERVOIR

PLASTIC TUBES

TAILING
CORE ~

RETORT
STAND-

WATER
LEVEL

^LEACHATE

LEACHING EXPERIMENTS— SETUP OF APPARATUS

FIGURE 3

- 544 -

Table 1 Chemical Characteristics of Leaching Solutions

Dilute

Unenr iched

6.0

Conductivity

4,5

(uohms/cm)

Total Phosphorus

<.01

(mq/l)

Nitrate

<.01

(ipg/l)

Ammonia

<.01

(mq/1)

Higher Strength
Enriched

350-400
1.5
0.86

3.45

Each column was dosed with solution until four or five
leachate fractions of about 100-150 ml each had been
collected. The average daily dose was in the order of
40 ml.

The treatments were ordered such that the expected most
severe conditions were appl ied last. The order of the
treatments was as follows,

1) Flush of residual pore water

2) Dilute, watertable below surface

3) Dilute, watertable above sur face, aerobic

4) Enriched, watertable below surface

5 ) Enriched , watertable above sur face, aerobic
r>) nilutf, watertable above sur face, anaerobic
7) Enriched, watertable above surface, anaerobic

After each treatment, the
followed on continuously.

cores were not drained; each

When the columns were collected, pore water was retained in
the tubes. As a result, the first series of samples
collected were in fact not a result of the treatment being
applied but were rather the discharge of the residual waters.
To deal with this lag, a final leaching solution was used
after the six treatments with a high concentration of
chloride (conductivity >3 ,000 uohms/cm) . Samples were
col lee ted and submitted for analysis from each column until
the chloride solution appeared . The leaching was terminated
at that point. Based on the volume of the fractions

- 545 -

r-o I 1 .m: t (.(1 bflwtM'n I lu' ond of Mio ^ .^r,\ h rea tmen l^ nnd lUv
appenrance of the chlocide solution, the leachates samples
were matched to the treatments by shifting the results by the
appropriate lag for each column.

Bioavailability

A primary concern wi th the leachate samples collected was the
proportion of heavy metals in solution that was biologically
available for uptake by plants and animals, A type of
bioassay procedure was used to provide an estimate of
bioavailabil ity .

Each leachate fraction was separated into two parts. One was
used as a growth med ium for duckweed ( Lemna minor ) and the
other was subm itted directly for analysiTT The part used to
test bioavailabil ity was supplemented (eg , 1 ml/100 ml
leachate) by a stock solution of nutrients required for
growth by duckweed. Then approximately 100 plants were
placed in 250 ml plastic beakers. Controls were grown in
distilled water supplemented by growth medium. The beakers
containing the duckweed were placed in growth chambers and
grown under controlled light and temperature conditions for
seven days.

All plants in each beaker were then harvested, rinsed with
distill ed water, dried and weighed . The dried samples were
d igested for 4 8 hours in a nitric acid solution. On
completi on of the d igestion, the samples were f il tered and
submitted for heavy metal analysis.

Al 1 samples were analyzed using an Inductively Coupled Plasma
Analyzer , Model Number 3400 manufactured by Appl ied Research
Laboratories,

RESULTS

A total of nine columns were used in the laboratory
experiments. They were collected from various tailings
deposits in the Cobal t area . Table 2 provides physical data
describing the strata of each column and Figure 2 indicates
the location and edaphic conditions of the sites at which
each column was collected. There are two columns from each
site except Site 4 for which there are three.

tabu: 2 suwARY OF mreiCAL dato for leaching couwre

Colirri Site Core Nunber
^^J-ber Nkjnber [ength of

(on) Strata

DESCRIPTION OF STRATA

1 103

■2

120

-4

2 72

J 80

■ft

U- '

U '

13 J

1
A

1 2

« 3
A

1 4

1 5

1 6

• 7

Leaching
Rate

Ott

S-M

o*t

nH-ft

Of ft

dct

riM-ft

O+ft

o*t

•

Oft

-F

5£

CODUJG:

I'JV

n Heart ings

Texture

Composition

Colour

Leaching Rate

- indicates strata sequence
in descending order
starting at surface

c^ coarse

f-fine

m-maJiim

o-organics

s-sand

t-tallings

d-dark
1- light

F-fast

M-moderate

S-slow

A
B

- Strata Length (on)

- physical characteristics
of strata

- 547 -

The trends in leachate chemistry from the columns were
considered at several levels, namely

i) differences among sites
ii) differences within sites
i i i) differences across treatments

Those trends were examined both in terms of hotn] dissolved
met a] concent rat ions and bioava J ] able fractions .

Differences Among Sites

piezometers were installed at each of the four sites that the
columns represent , Based on recorded groundwater
concentrations of heavy metals, the sites were ranked from
most to least contam inanted as shown.

Rank

Site Depth 0.9
1.5

Highest
1

Lowest
4

The sites were ranked based on piezometers at both the 0.9 m
and 1.5m levels.

A similar procedure was followed for the columns by grouping
them by site and averag ing contaminant concentration for the
columns in each group (Figure 4), The results were:

Highest

Lowest

Rank

Element

1,3

It can be seen that the relative ranks of the sites remained
similar to that for the 1,5 m depth piezometers.

Interestingly, the high concentrations in the shallow
groundwater of Site 1 are not seen in the leachate from the
columns. The columns from this site and, in particular.
Column K 2 hml low permeabilities and the lowor str.it.i
r(»ns i s I of n.i t. i vc^ s i 1 (■ y so i 1 s and a na tur ri 1 orf^an i c 1 ayo r ,
These layers appear to be i ntorcept i ntj part of the met.ils
being leached from the upper strata.

- S4S -

3
a.
a.

z:
o

z
o
u

2
ft.

z
o

bJ
U

z
o
o

AV!^::^AGE HF:AvY ^^^lAL CONCENTRATIONS

COLUMN tEACHATE WATER

0.6

0.5

0.4 -

0.3

0.2 -

0.1 H

E$3

7^ >M'A

v)^m^

Heavy Metal Mobilization
and Bioavailability
Cobalt Mine Tailings

J E Hanna
Associates inc.

FIGURE 4

- S49 -

Overall, the columns appear to have behaved similar to the in
situ tail ings ,

Dif ferences Within Sites

The response of each column to the treatments was examined
with respect to their representative sites . Some columns
tended to show the same trends, for example Columns # 1 and #
2, Columns # 4 and # 5, Columns # 12 and # 13. However
Columns # 7 and # 8 and Column # 11 compared to Columns # 12
and # 13 showed quite divergent responses. Where two columns
from the same site exhibited a similar response, the results
for only one column from that site is shown (Figures 5 to
10).

A large number of explanations could be postulated for these
differences between columns from one site but no single
explanation is comprehensive and convincing. The columns
from each of the sites appear to have similar physical
characteristics and no single parameter appears to correlate
to the observed d i f ferences . This variability draws
at tent ion to the com pi ex i ty of the geochem ist ry of tail ings
and in particular, of predicting heavy metal mobilization.

E ffects of Treatments

As discussed above, groups of the columns showed certain
common trends but no consistent trend was apparent among all
columns.

All columns except Column # 2, tended to have higher
concentrations of one or more metals in the residual water
than they did over most of the treatments. This may be due
to three factors :

i) The columns were sealed and stored for approximately 4
weeks before the experiments could begin. Over this
period, the residual pore water was able to reach
equilibrium concentrations,

ii) It is expected that anaerobic conditions increased in
the sealed tubes and redox potential dropped , increasing
the mobility of the metals.

i i i) Hydraul ic loads to the columns were much higher (ie, 10
to 20 times) than natural infiltration rates which may
have caused dilution compared to natural conditions.

Column # 7 showed no response to any of the treatments except
the first one and this is attributed to mixing of the
residual water with the incoming leaching solution and not a
change in leaching rate.

- 550 -

z
o

z
111
u
z
o
u

AVERAGE HEAVY METAL CONCENTRATIONS

LEACHATE: COLUMN 2

TREATMENTS

-t- Nl ■J

ARSENIC

'2
a.

*>^

z
o

5 ^

2 -

1 -

Heavy Metal Mobilization
and Bioavailability
Cobalt Mine Tailings

J E Hanna
Associates inc.

FIGURE 5

- 551 -

o.
ft.

z
o
p

hi
O

z
o
u

AVERAGE HEAVY METAL CONCENTRATIONS

0.45

LEACHATE: COLUMN 5

TREATMENTS

4- Nl

ARSENIC

2
a.
a.

z
o

z
o
u

Heavy Metal Mobilization
and Bioavailability
Cobalt Mine Tailings

J E Hanna
Associates /nc

FIGURE 6

- 552 -

AVERAGE HEAVY METAL. CONCENTRATIONS

3
0.
0.

-•^

z
o

"1
9 -

LEACHATE: COLUMN 7

' -—

8 -

7 -

6 -

■

5 -

■

4 -

3 ~

2 ~-

t -

T ¥ f

G CU

TREATMENTS

+ Nl C

ARSENIC

Heavy Metal Mobilization
and Bioavailability
Cobalt Mine Tailings

J E Hanna
Associates inc.

FIGURE 7

- 553 -

AVERAGE HEAVY METAL CONCENTRATIONS

a.
a.

Id
U

z
o
o

LEACHATE: COLUMN

0.11 -

0.1 -

0.09 '

O.OB

■r

0.07

0.06 -

t
0.05 -

"^^

0.04 ~

^^^^^v-^

---^

J«t^^^^

0.03 -
0.02 -

■ilh

\^^

■^

0.01 -

~r-

T ■ - - -■

TREATMENTS

ARSENIC

fL
0.

z
o

z
o
u

Heavy Metal Mobilization
and Bioavailability
..^'. Cobalt Mine Tailings

J E Hanna
Associates inc.

FIGURE 8

- 554 -

3
ft.
0.

'•—^

AVEf^AGE HEAVY METAL CONCENTRATIONS

LEACHATE: COLUMN 11

TREATMENTS

I NI

ARSENIC

2
a.
a.

Heavy Metal Mobilization
and Bioavailability
Cobalt Mine Tailings

J E Hanna
Associates inc.

FIGURE 9

- 555 -

2
a.
a.

z
o

z
o
o

AVERAGE HEAVY METAL CONCENTRATIONS

LEACHATE: COLUMN 13

TREATMENTS

1- Nl

ARSENIC

Heavy Metal Mobilization
and Bioavailability
Cobalt Mine Tailings

J E Hanna
Associates inc.

FIGURE 10

- S56 -

Column « 4 showed a response similar to Column # 7 for all
metals except arsenic which dec! ined steadily over the
experiment. A peak at Treatment # 1 may be caused by mixing
and initial short circuiting of the leaching solution but
this explanation is tenuous. Column # 11 showed a similar
pattern for arsenic with a peak at Treatment # 1 and a steady
decl ine until Treatment # 6.

Column H 'i 1 i kowi fic Followed this trend for rirsonic excopt
for a ma )or incre-ise wi th Treatment # 4 .

Columns # 1, # 2, # 8, # 11, # 12 and # 13 tend to show a
significant increase in the concentration of one or more
elements in the later treatments. Columns # 1, and # 2 show
a marked increase in arsenic wi th Treatment # 5 and zi nc
increases markedly wi th Treatments # 4 in Column # 1 and # 5
in Column # 2. Arsenic in Column # 8 peaked with Treatment #
5 and in Column # 5 with Treatment # 4,

In order to discriminate between treatments, it is assumed
that the leaching solution passes through the column in a
perfect plug flow pattern. This assumption does introduce
some error and the peaks at Treatment # 4 could reflect
partial ly the effects of Treatment # 5 and visa versa .
Regardless, it appears that Treatment # 5 and/or # 4 did
cause a major increase in the mobil ity of some elements, in
particular arsenic.

The dilute solution combined with the anaerobic conditions
was expected to increase leaching rates due to the higher
solubility of some elements under these conditions (Bolter
and Butz, 1973). Why this response only occurred in some
columns and not others cannot be determined at this time.

The treatments were sequenced in the expected order of
increasing severity wi th leaching rates expected to increase
generally from Treatment # 1 to Treatment # 6. Some elements
in some columns cooperated in following this trend, for
example zinc in Columns # 1, # 2, # 8, # 11 and # 13.
Arsenic responded in a similar way in Columns # 12 and # 13.
These two elements were the most responsive of the four to
the treatments. The interesting question in these results is
the absence of the overall expected trend in some columns and
the high variability between columns.

The results suggest that:

i) the chemical behaviour of heavy metals is difficult to
predict for all cases;
ii) anaerobic conditions can cause increased heavy metal
mobilization and
i i i ) low ionic strength solutions may cause increased
leaching .

- 557 -

Effects of Treatments on Bioavailable Fraction

The biomagnif ication ratio is defined as the concentration of
the element in the plant tissue divided by the concentration
in the solution (Woolson, 1975). Figure 11 illustrates the
average biomagni f ication ratios for duckweed for each of the
treatments- The values range from about 100 to 2600 times
the concentration in the leachate. These rates are in the
range of those reported by Clark et al (1981) for duckweed
for the metals considered.

The trends for the bioavailable fractions of arsenic and
copper closely matched that of their concentrations in the
leachate as a result their biomagni fi cation ratios are quite
constant. Nickel was quite variable compared to copper and
arsenic but no clear trend is apparent over the treatments.
Zinc, however, tends to behave quite differently under the
anaerobic conditions ( ie, residual waters and Treatments # 5
and # 6) . Overall , the bioavailabil ity of this element
appeared to increase with anaerobic conditions over the
proportionate concentration in the leachate waters. it may
be that this availability is due to different forms of the
metal in solution which make it more readily ingested by
duckweed .

Regardless of the specific mechanism, it appears that high
watertablG and anaerobic conditions tend to increase the
bioavailability of zinc but not arsenic and copper and they
may cause some increase with nickel.

CONCLUSIONS

Intact cores of tail ings behaved consistently with field
observations of contaminants concentrations . The
experimental leaching methodology developed in this study is
a reliable means to simulate the effects of chang ing
environmental conditions on tail ings geochemistry.

A high degree of variability is clearly apparent among the
tailings sites studied. The variability relates to obvious
physical features such as moisture and nutrient status,
vegetation, physical texture and age of tail ings. However ,
what appears visually to be the same material taken from the
same site may respond quite differently in terms of leachate
concentrations .

This variability demands that careful monitoring of any

activities (be they tailings management or marsh

construction) is undertaken in combination with normal
analysis, prediction and design procedures.

- 558 -

TISSUE/WATER CONCENTRATIONS
(Thousands)

o
o io

o
at

(30

— N) 4^

S . ,

: 1

o
c

—A ---.-«

N.
3

W
'^

. -i

\ \.

A.-

\ .\

:v■^v^x,^v^

' i

-n
O

o
o

i r

c
s:

ST"-

— ^.-w

ivi3SSr!?iS!SliSv.:::>S:':SSJ§S^ ! h £:

r v "^-r '■^.- T, v-^ \ \ \ X V X V ' - "^

Ol r^

"n]

m
z

-I
(/)

3--.

"T1

O
if)

Heavy Metal Mobilization
and Bioavailability
Cobalt Mine Tailings

J E Hanna
Associates inc.

FIGURE 11

- 559 -

3. Three factors considered in this project, namely i) ionic
strength of leaching solution and nutrient availability,
ii) soil moisture status, and iii) aerobic vs anaerobic
environments, all had noticeable effects , either individually
or collectively, on heavy metal mobil ization. Anaerobic
conditions had the greatest effect followed by ionic strength
of the leaching solution. Watectable conditions had the
1 cist ef feet .

However, high leaching rates of metals occurred under all of
the chemical and physical environments used in the treatments
in one or more of the columns. Accordingly, heavy metal
leaching is a concern with the Cobalt tailings regardless of
the chem ical and physical environment, but, the greatest
potential exists with elevated watertables and anaerobic
conditions.

4. The use of duckweed provided reasonable biomagnif ication
estimates and consistent results. The technique was
relatively simple and gives a direct estimate of the
bioavail able fraction.

'j . The bioavailable fraction of metals in the leachate was
directly proportional to the concentration of the metals in
solution and remained relatively constant over the treatments
tested except where an anaerobic env ironment ex is ted . This
observation suggests that the form of the metals in solution
changes under these conditions.

6. Bi ©magnification of heavy metals in pr imary producers ( ie,
duckweed) was in the order of 10 to 2600 times that in
solution. Many of the plants grown in the leachate solution
showed serious toxic stress from the contaminant levels and
accumulation rates at higher trophic levels may be greater.
Accordingly leaching of heavy metals from tailings presents a
significant toxic hazard to biological organisms in the area.

7 , The pr imary concerns relating to increased metal leaching
from the construction of a marsh treatment system are listed
in order of importance:

a) creation of strong anaerobic conditions

b) development of elevated watertables, greater
leachate quantities and more rapid groundwater
velocities

c) physical disruption to tailings and exposure of new
material to leaching process

d) supply of organic acids to adsorb metals

e) increase biological uptake and availability through
enhanced primary production.

- 560 -

BIBLIOGRAPHY

Allen, n.E., and M.T, linger, 1980. Evaluation of Potential
Metal Mobil ization from Aquatic Sed iments by Complex ing
Agents. Z. Wasser, abw. Forsching, 13. p. 124-129.

Bolter, E. , and Butz, T.R. 1976, Heavy Metal Mobilization
by Natural Organic Acids, Proc. of the International
Heavy Metals Conference 1975. Toronto, Canada. p. 353-
362.

Clark, J.R., Vanhassel, J.H., Nicholson R.B., Cherry D.S.,

and Cairns J. Jr. 1981. Accumulation and Depuration of

Metal s by Duckweed . Ecotox icology and Env ironmental
Safety 5: 87-96.

Down , C.G. , and Stocks . j, 1978 , Environmental Impact of
Mining. Applied Science Publishers Ltd., London,
England ,

JE Hanna Associates Inc. 1984. Heavy Metal Mobil ization and
Bioavailability - Cobalt Mine Tailings. Draft report
submitted October 1984 to Ontario Ministry of the
Environment. p. 115 + app.

Jenne, E.A., and Luoma, S, No date. Forms of Trace Elements
in Soil s, Sed iments, and Associated Waters: An Overview
of Their Determination and Availability. U.S.
Geological Survey, Menlo Park, California.

Peters, T.H. 1978. Inco Metals Reclamation Program in Proc .
Second Annual Meeting of the Canad ian Land Reclamation
Association. Laurentian University. May-June -78.
ISSN-0705-5927.

Wool son, E.A. 1975. Bioaccumulation of Arsenicals. Chap.
7. In: Arsenical Pesticides. Edited by E.A. Woolson.
ACS Symp. Ser 7. p. 97-107,

- B61 -

Water quality analysis of trout farm effluents

J.W. Hilton, G. Chapman and S.J. Slinger

Department of Nutrition

College of Biological Science

University of Guelph

Guelph , Ontario

NIG 2W1

- 562 -

WATFR OIJAI ITY fHARACTFRISTICS OF TROUT FARM EFFLUENT
J.W.Hilton , G. Chapman and 5. J. Slinger

Department of Nutrition
College of Biological Science
University of Guelph

Guelph Ontario NIG 2WI
The water quality characteristics of eight private trout farms in
Ontario were investigated over a four month period from June to October
1983. Each farm was visited three times during this period and parameters
such as water flow rate, trout farm biomass .feed and feeding system and
water chemical parameters such as ammonia CNH3), and phosphorus
(P, total phosphorus) were measured and collected. The results indicated
that the major water qulaity parameters affected by the trout farms were
NM3 and P. The daily loading of P from ttie trout farms ranged from 59 to
1299 g/day which was highly correlated to the biomass of the individual
trout farm. In addition, the NH3 loading from the trout farms ranged from
1 8 to 141 kg/day. However, in contrast to the P loading, the NH3 loading
did not appear to be highly correlated to the trout farm biomass which was
unusual considering that ammonia is supposed to be the major nitrogenous
waste (>90%) of fish such as rainbow trout.There appeared to be some
beneficial effect of having either a settling or a retention pond at the
trout farm in order to reduce the loading of NH3 or P. However, this
conclusion requires further study and verification due to lack of
standardization of the various trout farms. In addition to the trout farm
effluent study, a number of diet-growth studies were conducted to
determine the effect of diet on phosphorus retention in rainbow trout.
These studies mdtcated that supplementation of commercial trout diets.
containing 25% fish meal in the feed formulation, with dicalcium phophate
was unnecessary. Furthermore, the formulation and feeding of high
protein^energy diets to rainbow trout significantly increased phosphorus
retention in these fish On the basis of these results , it would appear to
be possible to reduce phosphorus excretion in the trout by dietary
manipulation. The formulation and processing of low-pollution trout diets
could be of benefit to trout farms were NH3 and P loading exceeds
governmental regulations.(Supported by OMAF and OME)

- S6.S -

Introduction

There is a growing awareness that trout farms can be
potential polluters of receiving waters. Hinshaw (1973) was
perhaps the first to indicate that trout farms can cause a
degradation of water quality downstream from the trout farm
effluent input. The major factors to which he attributed these
changes were the size of the farm and the volume of the receiving
waters. Surprisingly, Hinshaw states that there is no
correlation between the type and quantity of food fed at the
trout farm and the changes and/or degradation of the water
quality of the receiving waters. However, he did no experiments
relative to diet or feeding. Obviously the source of the
limiting nutrients such as nitrogen and phosphorus must
ultimately come from the food that fish consume, and therefore it
is difficult to accept Hinshaw's conclusion. Although Hinshaw
did measure several water quality chemical parameters such as
dissolved oxygen (DO), ammonia (NH^), nitrate (NO^) and nitrite
(NOp), he relied primarily upon the changes in the population and
occurrence of bottom fauna. Furthermore, while he did sample
water from a number of different farms, he did not sample over a
24 h period and did not usually have any repeat visits to the
farms. In contrast, Tervert (1981) conducted a detailed study of
one particular trout farm (as well as a survey of other farms)
over the period of a year (July 1975 to February 1977). The
major parameters he measured were biological oxygen demand (BOD) ,
DO, NH-, suspended solids (SS) and total phosphorus (TP). On the
basis of his results, Tervert developed a formula which generated
a water quality index (WQI). According to Tervert, this

- 564 -

particular index could be used to ascertain the extent of water
quality degradation of the receiving water by a trout farm. The
major drawback of the WQI is that it requires that each farm be
examined at frequent intervals in order to estimate the mean WQI.
Furthermore, variations in the flow rate of the receiving waters
must also be taken into account. While these measurements may be
desirable, their determination may not be practical for most
trout farms. Perhaps the most extensive work done to date, on
the water quality characterstics of trout farm effluents is that
of Bergheim and coworkers (Bergheim and Selmer-Olsen , 1978;
Bergheim et al , 1982, 198M), These studies have concentrated not
only on a single high density trout farm but also a number of
small trout farms in Norway. Generally their studies have been
conducted for over a year, during which time the farms were
visited several times and with repeated sampling times on each
visit. Furthermore, their measurements were much more extensive
than either Hinshaw's or Tervert's. Nevertheless, as in the
other studies, the major parameters that were affected by the
trout farm operation were BOD, NH- , NO-,, TP, SS and total
nitrogen (TN). Despite the fact that Bergheim* s group went to
the trouble of collecting additional data on the size, number and
species of fish, as well as type of food and feeding system, they
did not relate these factors to the major water quality
parameters that they measured. Furthermore, only one of the
farms had a really high population density.

Aside from the biomass and water flow rate of the trout
farm, the other major factors which could affect the water

- 56S -

quality characteristics of the trout farm effluent are the
composition and pellet durability of the fish feed, its
phosphorus content and availability, the feeding system used by
the farmer, the water temperature and the presence or absence of
either settling or retention ponds. It would seem likely that
some prediction on the potential loading of receiving waters may
be obtained by relating all the diverse factors. Furthermore,
since the fish food itself is the ultimate source of the waste
materials and nutrients that are being voided into the receiving
waters, research on the development of low pollution diets for
fish could prove to be very advantageous in terms of reducing
effluent loading of receiving waters.

The purpose of this study was to determine and/or
investigate :

1) the effect of seven different private and two public trout
farms on the water quality characteristics of the farm
effluent.

2) the necessity of supplemental phosphorus inclusion in the
formulation of commercial trout diets in Ontario.

3) the effect of dietary manipulation on the phosphorus
retention and excretion in rainbow trout.

A. Trout Farm Effluent Study

Experiment Design - Materials and Methods

During the summer and early fall of 1983 seven commercial
trout farms in Southern Ontario were visited three times. The
visitation times were selected to cover the peak production
(highest biomass) periods of the trout farms in Ontario. At the

- r>()6 -
time of each site visit, the individual characteristics of each
farm were noted such as: number of raceways; ponds and/or tanks;
estimated biomass; feeding system; type of feed; presence or
absence of settling ponds and retention ponds; and size
distribution of fish. Water samples were collected at a number
of different locations throughout the trout farm. These water
samples were collected at three different time periods during the
visit. The time spent at each farm during each visit was
approximately 7-9 hours. Normally, water samples would be
collected in the morning, at noon and around 4 to 5 o'clock in
the afternoon. At each location a number of different water
parameters were measured at the same time water samples were
collected. These parameters included: water flow rate (velocity
meter), pH (Orion pH meter), DO (EYI dissolved oxygen probe), NH^
(Nessler*s reagent) and water temperature. Water samples were
analyzed in the Department of Nutrition, University of Guelph
for total phosphorus (Persulphate - ascorbic acid method,
Standard Methods 1980). An attempt was made to monitor the
impact of raceway cleaning on water quality parameters. However,
few of the trout farms were cleaning during the site visits,
therefore this information is discussed only as it pertains to
individual fish farms. In addition to the private fish farms,
two public hatcheries were also included in the survey. These
included Chatsworth Hatchery and the North Bay Hatchery, both
operated by the Ontario Ministry of Natural Resources (OMNR). It
should be noted that while the Chatsworth Hatchery was sampled by
Hilton and coworkers, the North Bay Hatchery was sampled by OMNR
personnel .

- 567 -

B. Trout Feeding Study

Experimental Design

Experiment I - Essentiality of supplemental inorganic phosphorus

in practical trout diets.

Juvenile rainbow trout were reared for 12 weeks on
3, low-fish meal content practical trout diets
supplemented with 0, 1 and 2% dicalcium phosphate
in a randomized block design.

Experiment II - Phosphorus retention study

Juvenile rainbow trout were reared on U,
practical-type diets for 16 weeks. The U diets
were a control standard trout diet and three test
diets formulated to have higher protein and lipid
levels than the control diet, but the same level
of dietary phosphorus. The study was conducted
using a randomized block design.

Diet formulation and processing
Experiment I

Three test diets were formulated as outlined in Table 11,
The test diets were processed by steam pelleting on a laboratory
pellet mill and then stored in a cooler until required for
feeding. The diets were analyzed after processing for ash, crude
protein, and moisture content as described by Horwitz (1980), and
phosphorus content by atomic absorption spectrophotometry.
Experiment II

Four diets were formulated as outlined in Table 12 with diet
1 as the control diet. The test diets were processed by steam

- S68 -

TABLE

Formulation, proximate composition

and

phosphorus content

the trout

diets in experiment

X (phosphorus supplementation study)

Ingredient

Diet Number
2

(%)

Capelln Meal

Soybean Meal

Wheat Middlings

Brewer's Yeast

Com Gluten Meal

Poultry By-Product Meal

Alfalfa Meal

Vitamin Mix

Mineral Mix

Calcium-Phosphate

Kaolin

Fish Oil

Analysis-Protein

37.5

37.4

36.9

Lipid

18.3

18.7

18,9

Ash

7.8

7.7

7.4

Phosphorus

1.4

1.7

2.0

- S69 -

TABLE 12

Formulation, proximate composition and phosphorus content of
trout diets in experiment II (high protein:energy diets)

Ingredient

Diet

Number

Capelin Meal

Blood Meal

■^^

Poultry By-Product Meal

—

Feather Meal

—

Com Gluten Meal

— .

—

Wheat Gluten

—

Soybean Meal

—

■_-_

Wheat Middlings

—

— —

■■"

Vitamin Premlx

Mineral Premlx

Bentonite

—

Fish Oil

Analysis-Protein

39.8

58.1

56.3

58.2

Lipid

17.0

24.9

24.8

26.2

Ash

8.8

9.2

13.2

12.5

Phosphorus

1.6

1.3

1.5

1.7

- 570 -
pelleting on a laboratory pellet mill, and then stored in a
cooler C-5°C) until required for feeding. The diets were
analyzed after processing for ash, crude protein and moisture
content as described by Horwitz (1980), lipid content by the
method of Bligh and Dyer (1959), and phoshorus content by atomic
absorption spectrophotometry .
Supply and maintenance of fish

The test diets were fed to either triplicate groups in
experiment I or quadruplicate groups in experiment II of 70
juvenile rainbow trout (initial weight 4.9+0.2 g/fish) for 16
weeks. The trout were obtained from a commercial trout farm and
acclimated in the laboratory for approximately 2 weeks prior to
initiation of the experiment. The trout were maintained in
rectangular fiberglass tanks (volume 60 L) that were individually
aerated and had a water flow rate of approximately 2 L per min.
Water temperature was thermostatically maintained at 15.U±0.3 C
and the dissolved oxygen and pH were monitored weekly and ranged
from 7.1 to 8.1 mg/1 and 7.8 to 8.0, respectively, throughout the
test period. The tanks were housed in a windowless laboratory
which had a photoperiod of 12 h light and dark supplied by
fluorescent lighting. The trout were fed three to six times
daily to satiety as described by Hilton and Slinger (1981). The
trout were weighed at the end of each 28-day period, and the size
of feed particle adjusted after each period. Mortalities were
monitored daily and feed:gain ratios determined after each
period .
Biochemical analysis

After 12 or 16 weeks on the test diets, approximately six

- S71 -

fish were removed at random from each tank, anaesthetized with
tricaine methanesulphonate (MS222) and blood collected by
amputation of the caudal penduncle. The hemoglobin content of
the blood was determined by the cyanmethomeglobin method and
haematocrit levels by microhematocrit tubes. The fish were then
euthanized by severing the spinal cord behind the head. The fish
were then ground in a meat grinder, frozen, freeze-dried and
analyzed for dry matter, crude protein, lipid, ash and phosphorus
content as previously described.
Statistical analysis

The data were subjected to analysis of variance and
treatment significance determined at the 5% level using Tukey* s
honestly significant difference procedure as described by Steel
and Torrie (1980) .

Results - Trout Farm Effluent Study

Trout Farm A - Springhills Trout Farm, Chatsworth, Ontario

Visitation Dates - June 28, August 9, September 20, 1983

General Description

Water Supply and Flow Rate - The farm is supplied by three main

springs with a total flow rate of 1539 to 2700 L/min during

the site visits.
Buildings and Raceways - The farm consists of H greenhouse type

buildings each containing three cement raceways and

additional fiberglass tanks for a total of 12 raceways and

10 fiberglass tanks.
Biomass - The estimated biomass ranged from 6U50 kg in June to

12500 kg of fish in September. The farm was not at full

- ^12 -

TABLE 1
General water quality parameter measurements - Trout Farm A

Watt

er-Sample Sit*

Flow rate (L/mln)

1589-2700

N.D.

1589-2700

Temperature ('

•c)

9.4-10.0

9.1-10.0

9,1-10.1

9.4-11.4

DO (mg/L)

8.0-12.0

8.0-10.1

7.3-9.0

7.0-9.4

6.9-9.5

7.53-7.61

7.4-7.58

7.51-7.78

7.4-7.8

7.42-7.81

NH^ (mg/L)

<.l

0.3-0.9

0.35-1.2

1.2-1.5

0.8-1.52

TP (yg/L)

B
C

6-12
6-15

12-49

12-35

34-95

40-111

34-114

^Results are expressed as the mean of the 3 water samples collected per visit
and Klven as the range from the lowest to the liighest readings for the 3 visits

2
Not determined.

Different spring phosphorus readings.

Water-Sampling Sites

A - All three main springs prior to entry into the farm
B - End ot the first greenhouse terminal raceway
C - End of the last greenhouse terminal raceway
D - Prior to settling pond
E - Exit from settling pond

- S77> -

capacity.

Feeding System - With the exception of the brood stock and swim-
up fry to juvenile fish, all fish were fed a sinking pellet
by way of demand feeds. The remainder of the fish were fed
manually.

Settling Pond - A cement settling pond was connected to the end
greenhouse and terminal raceway with all water passing
through this settling pond prior to leaving the farm.

Retention Pond - The retention pond was connected to the settling
pond and was not used during any of the site visits.
However, a drain in the settling pond was connected via a
pipe to the retention pond. The retention pond had no
visual and/or apparent outflow.

Trout Farm B - Aberfoyle Fisheries, Aberfoyle, Ontario

Visitation Dates - June 29, August 10, September 21, 1983

General Description

Water Supply and Flow Rate - This farm is primarily supplied by

well water with a flow rate of 2500 to 2719 L/min. However,
this farm also uses a recirculation system with
approximately 30-40^ of the water recirculated through
gravel bed ponds prior to reuse.

Buildings and Raceways - The farm had one large building which
housed approximately 57 fiberglass tanks. The broodstock,
larval and juvenile fish were maintained in this building.
Outside the building were 18 large cement raceways with
three raceways linked in series by gravity water supply. It
should be noted that this farm employed aeration systems

TABLE 2
General water quality parameters - Trout Farm B

Parameter

Water-Sample

Site

Flow rate (L/min)

N.D.

2501-2719

325-378

Temperature (*'C)

8.0-9.8

9-9.5

9.8-10.2

10.3-11.5

10.2-14.5

10.2-14.8

11.2-17.0

DO (mg/L)

4.8-8.1

7.5-9.2

6.8-8.9

5.6-6.4

4,2-7.6

4.5-7.0

0.5-2.2

7.0-7.45

7.1-7.46

7.3-7.5

7.1-7.3

6.9-7.35

6.9-7.4

7.0-7.14

N-H^ (mg/L)

0.25-1.10

0.7-0.9

0.8-1.4

2.4-2.8

2.6-3.5

2.8-3.7

0.8-2.6

TP (ug/L)

5-54

35-42

38-44

245-369

178-301

194-345

889-1513

"Results are expressed as the mean of the 3 water samples collected per visit and given as the
range from the lowest to the highest readings for the 3 visits.

"Not determined.

Water-Sampling Sites

A - Inside the building - mixture of well and recirculated water

B - End terminal raceway inside building

C - Beginning of 3 raceway systems outside of building

D - End of 3 raceway systems outside of building

E - Prior to entry into settling pond

F - Exit from settling pond

G - Exit from retention pond

- 575 -

with injected oxygen into the outside raceways.
Biomass - The farm had the largest estimated biomass throughout

the study ranging from 53,011 kg to 57,705 kg of fish. The

farm 'appeared' to be near full capacity.
Feeding System - The farm relied on manual type feeding systems

inside the building and demand feeding systems outside the

building. Type of fish food was a sinking trout pellet.
Settling Pond - The farm had one main settling pond which

connected either to the creek or with a recirculation pond.
Retention Pond - The farm had two retention ponds which were

connected in series. The second retention pond drained by

way of a pipe into the nearby woods. During all three

visits this pipe had water continuously flowing into the

woods.

Trout Farm D - Franklin Trout Farm, Mount Albert, Ontario

Visitation Dates - July ^1 , August 15, September 26, 1983

General Description

Water Supply and Flow Rate - This farm is supplied by both spring
(well) and lake (pond) water. Spring water supplied
primarily the broodstock and post-larval fish at a flow rate
varying from 605-1125 L/min. Pond water supplies the
juvenile and grow-out trout at flow rate of 4996 to 5170
L/min. It should be noted that the spring water augments
the pond water supply as a means of controlling (partially)
water temperature .

Buildings and Raceways - There are no buildings housing any

raceways or tanks on this farm. The farm has 3 concrete

- S76 -

TABLE 3
General water quality parameters - Trout Farm D

Parameter

Water-Sample Site

Flow rate (L/mln)

605-

-1125

N.D.^

A996-5170

5327-

-6161

Temperature ('

^C)

9.2-

-10.5

9.9-14.2

12.0-19.2

12.2-

-19.9

(mg/L)

6.5-

•10.2

7.7-10.2

5.8-10.4

8.5-

-11.6

7.74-

-8.7

7.7-8.7

7.77-8.5

7.7-

-8.7

NH^

, (mg/L)

0.2-

-0.5

0.9-1.25

0.4-1.1

0.7-

-1.3

(lig/L)

16-

■19

82-95

33-76

39-

-112

Results are expressed as the mean of the 3 water samples
collected per visit and given as the range from the lowest to
the highest readings for the 3 visits.

2
Not determined.

Water-Sampling ^ _^A^.P.^

A - Spring water inflow

B - End of concrete raceway - terminal spring water

C - Inflow pond (+ spring) water

D - Outflow terminal grow-out circular raceways

- 577 -

raceways, 6 large circular tanks and 28 circular raceways.

Biomass - The initial biomass during the first visit was

estimated to be 26,775 kg of fish. However, by the time of
the last visit the biomass had been reduced to 9545 kg. The
initial biomass was certainly below the potential capacity
of this farm considering its water flow rate.

Feeding System - The farm relied upon manual feeding of

broodstock and post-larval to juvenile trout. However, the
majority of the farm juvenile to grow-out trout were on
demand feeders. The farm used primarily sinking trout
pellets (steam pelleted).

Settling Pond - none

Retention Pond - none

Tr out Farm E - Aquafarms Ltd. - Feversham, Ontario

Visitation Dates - July 5, August 16, September 27, 1983

General Description

Water Supply and Flow Rate - This farm is supplied by both spring
water and river water at proportions. The river water is
pumped to a header raceway. and then mixed with the spring
water at a ratio of approximately 15% river:25% spring
water. The water flow rate varied from 8768 to 10847 L/min
during the visitation periods.

Buildings and Raceways - One building housing post-larval and
juvenile fish in rectangular fiberglass tanks is located
some distance from the major raceway systems. This facility
was not sampled during this study. The major portion of the
farm consists of 10 plastic lined raceways. One of the

•11

- S7S -

TABLE 4
General water quality parameters - Trout Farm E

Parameter

Water-Sample Site

Flow rate (L/min)

Temperature ("C)

DO (mg/L)

NH^ (mg/L)

TP (Mg/L)

8768-10847

10.9-14.2
8.5-10.8

7.29-8.7
0.2-0.3
7-61

N.D.

11.5-17.2

6.2-8.6

7.5-8.13

0.4-1.2

16-40

8849-10847

11.3-17.0

7.2-8.7

7.66-8.7
0.48-1.2
20-74

Results are expressed as the mean of the 3 water samples
collected per visit and given as the range from the lowest
to the highest readings for the 3 visits.

Not determined.

Water-Sampling Sites

A - Inflow water mixture of river and spring water in
header raceway

B - Terminal raceway of one series

C - Major outflow of the farm

- 579 -

raceways lies perpendicular to the other raceways and
functions as a header for the remaining raceways. These
raceways are lined up in a series of 3 systems each
containing 3 raceways,

Biomass - The biomass of the farm ranged from 25000 kg to 3^090
kg of trout over the study period.

Feeding System - This farm utilizes an automatic feeding system

in which feed is blown by a hydraulic pump into each raceway
at discrete time intervals throughout the day. The type of
feed utilized is a sinking steam pelleted diet.

Settling Pond - None

Retention Pond - The design of these raceways results in the

accumulation of waste at the bottom of each raceway. This
waste is vacuumed into a retention pond (?) - structure
which has no outflow pipe. The owner-operator states that
the waste is removed from this structure and used as
fertilizer in other agricultural areas.

Trout Farm F - Blue Spring Trout Farm - Hanover, Ontario

Visitation Dates - July 6, August 17 and September 28, 1983

General Description

Water Supply and Flow Rate - The only water supply of this farm
is a spring which flows by way of gravity into the farm.
The flow rate ranged from 3^01 to 5170 L/min during the
study. It should be noted that this farmer employs a
biological-mechanical type filter to recondition his water.
This filtration apparatus is located approximately half-way
between the two major raceway systems on the farm. The

- BSO -

TABLE 5
General water quality parameters - Trout Farm F

Parameter ''' Water-Sample Site

A B C D E

Flow rate (L/min) 3401-5171 N.D.^ N.D, N.D. 3349-5217

Temperature (''C) 8.8-9.5 8.9-10.9 8.8-11.0 9.8-13.0 10.2-13.7

DO (mg/L) 9.6-11.2 7.8-10.2 5.1-8.5 7.2-10.6 5.5-9.8

pH 7.31-7.58 7.4-7.7 7.6-7.8 7.5-7.7 7.4-7.8

NH (mg/L) 0.25-0.40 1.2-1.6 1.2-1.6 1.7-1.9 1.5-1.8

TP (pg/L) 7-35 7-37 AO-71 29-104 35-81

^Results are expressed as the mean of the 3 water samples collected per
visit and given as the range from the lowest to the highest readings for
the 3 visits.

2
Not determined.

Water-Sampling Sites

A - Inflow water from Blue Springs creek

B - End of first raceway systems prior to filter

C - After filtration system

D - Prior to settling pond

E - After settling pond

- 581 -

farmer uses the initial spring water to supply the first
half of the farm, then filters and reuses the water to
supply the remainder of the farm.

Buildings and Raceways - This farm has no aquatic systems housed
in a building. There is a total of 30 raceways, 10 of which
are concrete and 20 of which are gravel-earth ponds.

Riomass - The biomass was estimated at around 16,500 kg which is
well below the maximum capacity of this farm.

Feeding System - All feeding was carried out manually during this
study, however, the farmer does have automatic feeding
systems in place. The feed was a sinking-steam pelleted
diet.

Settling Pond - One large earthen pond was used as a settling
pond prior to entry into the South Saugeen river.

Retention Pond - The farm also had a retention pond with a very
small outflow creek-pipe into the receiving water. Very
little flow was observed from this pond during the study.

Trout Farm G - Shamrock Springs Trout Farm - Erin, Ontario

Visitation Dates - July 7, August 18, September 29, 1983

General Description

Water Supply and Flow Rate - This farm is supplied by 3 main

springs and a number of smaller springs. In addition the
aquafer is very close to the surface such that some of the
gravel-mud ponds are drained through this aquafer. As a
result it is very difficult to accurately determine the flow
rate of this farm. On the basis of the three main springs,
the water flow ringed from 1556 L/min to 2266 L/min during

- S82 -

TABLE 6
General water quality parameters - Trout Farm G

Parameter Water-Sample Site

A B C D

Flow rate (L/miii) 1556-2266 N.D. N.D. N.D.

Temperature (°C) 8.2-9.8 10.2-lA.O 10.1-13.2 10.3-13,8

DO (mg/L) 9.4-11.8 9.7-10.5 9.2-11.0 8.5-12.4

pH 7.38-7.90 7.75-8.10 7.68-8.20 7.56-8.20

NH (mg/L) .1-.45 0.6-1.6 0.40-0.90 0.65-1.1

TP (ug/L) 0-17 15-140 15-171 6-82

Results are expressed as the mean of the 3 water samples collected
per visit and given as the range from the lowest to the highest
readings for the 3 visits.

2
Not determined.

Water-Sampling Sites

A - Incoming water to ponds (combination of 3 major springs)

B - Outflow from pond

C - Outflow from hatchery-laboratory

D - Major outflow from ponds (semi-settling pond)

- sss -

the study. However, this is probably an underestimate of

the actual water flow rate.
Buildings and Raceways - The farm has 8 concrete raceways, 6

earth ponds and a hatchery. The hatchery also houses a

small laboratory which contains a number of aquatic sytems.
Biomass - The total biomass of this farm ranged from 6600 kg to

4818 kg during the study.
Feeding System - The farm uses primarily a manual feeding system

with a sinking-steam pellet. However, the hatchery runs on

an automatic feeding system.
Settling Pond - There is no separate settling pond, however, on

the earth ponds does function as such for the farm.
Retention Pond - None

Tr out Farm H - Spring Valley Trout Farm - Petersburg, Ontario

Visitation Dates - July 8, August 19, September 30, 198M

General Description

Water Supply and Flow Rate - The water source is a combination of
approximately three springs and pond water. Spring water
primarily supplies the main hatchery and circular tanks.
Both pond and spring water are mixed together to supply a
series of raceways outside the hatchery. In addition, the
pond water supplies a completely separate series of
raceways. Both the pond and spring water mixture and the
pond water by itself are mixed altogether prior to exit from
the trout farm. The total water flow ranged from 9296 L/min
to 13031 L/min during the study period.

- 584 -

TABLE 7
General water quality parameters - Trout Farm 11

Parameter'

Water-Sample Site
BCD

Flow rate (L/min)

Temperature ("C)

DO (mg/L)

NH^ (mg/L)

TP (ug/L)

7.8-11.8
7.8-9.9
7.A3-8.1
0.1-0.2
0-17

N.D.

12.9-18.1

12.9-15.2

7.61-7.87

0.2-0.75

0-15

N.D.

12.4-16.4

6.1-9.8

7.6-8.0

0.7-1.6

51-459

N.D.

12.4-16.6

4.5-8.2

7.5-8.1

0.9-1.6

54-368

9296-13031
13.1-18.3

4.5-8.2
7.55-8.10
0.92-1.6
54-368

Results are expressed as the mean of the 3 water samples collected per visit
and given as the range from the lowest to the highest readings for the 3 visits,

"Not determined.

Impossible to accurately determine.

Water-Sampling Sites

A - Spring water inflow to hatchery (combination)

B - Pond water inflow to external raceway

C - Water outflow from hatchery

D - Water outflow (pond) from external raceways

E - Combined outflow water from trout farm

- 585 -

Buildings and Raceways - The farm has an enclosed hatchery which
includes 4 circular tanks and a number of smaller circular
tanks. There is a total of 8 cement raceways.

Biomass - The estimated biomass during this study was
approximately 20,000 to 30,000 kg of fish.

Feeding System - AH feeding at this farm was performed manually
with sinking-steam pellets.

Settling Pond - None

Retention Pond - None

Ontario Ministry of Natural Resources - Hatcheries

In addition to the private hatcheries, two public hatcheries
were included in the survey. These were the Chatsworth Hatchery
and the North Bay Hatchery. Although an attempt was made to
collect the same data as was collected for the private
hatcheries, it soon became evident that the public and private
hatcheries are not really comparable. The biomass in the public
hatcheries was so much less than that of the private hatcheries
that for many of the readings, there was very little change from
inflow. Furthermore, especially in the case of Chatsworth, the
flow rates were so much higher than that of the private
hatcheries, that again the water parameters did not really change
that much. The following table lists the basic characteristics
of the public hatcheries. Note that the results are a mean of
three sampling periods. It should also be noted that neither the
Chatsworth nor the North Bay Hatchery have either settling or
retention ponds. Furthermore, both hatcheries rely on manual
feeding systems using sinking-steam pellets.

TABLE 8
General water quality parameters of the Chatsworth and North Bay (OMNR) Hatcheries

Hatchery

Biomass
(kg)

Water Flow
(L/min)

Inflow

Outflow

Temp

DO
(mg/L)

NH3
(mg/L)

(yg/L)

Temp

DO
(mg/L)

NH3
(mg/L)

(yg/L)

Chatsworth
North Bay

800-1600
400-1200

13,503
2,250

7.5-10.2
7.5-9.0

9.4-10.4
9.6-10.4

0.1-0.2
0.1-0.2

17-43
25-84

7.5-10.0
7.7-9.0

6.8-9.0
8.5-9.0

0.1-0.3
0.1-0.2

15-42
38-80

- S87 -

Discussion - Trout Fartn Effluent Study

The results of this study-survey indicate that a trout farm
can significantly alter the water quality parameters of its
effluent and therefore could potentially cause considerable
alteration in the water quality of the receiving waters. This is
in essential agreement with the results of Bergheim and coworkers
(Bergheim and Selmer-Olson , 1978; Bergheim et al., 1982, 198U).
The extent to which the effluent could affect the water quality
of the receiving waters is beyond the scope of this study, but
obviously would be affected by such factors as the size and flow-
rate of both the effluent and receiving waters. Furthermore, due
to differences in size of the biomass, design and flow-rate of
the farms, the loadings from these farms did vary considerably.
Calculating the daily loading of either ammonia or phosphorus
from the trout farms (Table 9), indicates that ammonia input
varied from 1.8 to 14.1 kg/day and total phosphorus from 59-2255
g/day. Faure (1977) indicated that the waste-excretion from 1
kilo of trout is equal to that from 0.2 to 0.5 persons. Assuming
that this calculation is correct, some of the trout farms
examined in this study have the equivalent pollution impact of a
town of approximately 20,000 people.

The major water quality parameters affected by the trout
farms in this study were ammonia (NH^) , dissolved oxygen (DO) ,
total phosphorus (TP) and water temperature (Tables 1 to 8) . In
contrast to other studies (Bergheim et al., 1982, 198U),
suspended solids (SS) were not found to be significantly affected
in this study and were not listed in the results. However, the
previous study did include a 2U-M8 h sampling period which

- .S88 -

covered a complete cleaning of the farm. Only during the
cleaning periods were the SS elevated in the Bergheim studies.
In addition, although not indicated in the tables, there was no
variation in the effluent water quality parameters during the
sampling period. Bergheim et al. (1982, 1984) also noted no
changes in the water quality parameters of the trout farms except
during the cleaning periods when a dramatic increase in NH^, SS
and TP levels was noted. It should be noted that in the Bergheim
studies, the farms investigated did not usually have either a
retention or settling pond. It would be interesting to determine
the advantages of such systems in reducing the increased loadings
during cleaning from trout farms. Although it is very difficult
and perhaps unfair to directly compare trout farms in this study,
some interesting information may be obtained by comparing the
daily loadings of trout farms B and H (Tables 9 and 10). These
farms had the largest biomass of the trout farms visited, however
trout farm B had both a retention and a settling pond while trout
farm H had neither. The addition of both a retention and
settling pond did appear to reduce the loading of the trout farm.
However, further studies are required to validate these findings.
In addition, it should be noted that in the case of trout farm B,
the retention pond had an outlet through which the water would
eventually flow into the receiving water. This obviously should
not occur with a 'true' retention pond. One other trout farm in
this study had a retention pond, trout farm E; however, water
samples were not taken during the time when this retention pond
was being used. Nevertheless, it did not have an outlet pipe

- 589 -

TABLE 9

Total ammonia and total phosphorus output
per day from the trout farms

2
Trout Farm

Ammonia
(kg/day)

Phosphorus
(g/day)

3.1

147

10.3 (12.

.7)^

901 (2255)

3.7

343

8.0

302

5.3

229

1.8

14.1

1299

Results based upon the mean of the average dally
output of the farms during the 3 visits.

Does not Include OMNR hatcheries.

Retention pond outflow loading calculated with total

- S90 -

TAiU.K 10

Comparison of daily loading per unit blomass of
trout farms B and H

Trout Farm Biomass Total Phosphorus (g) Total Nitrogen (mg)
(kg) Biomass (kg) Biomass (kg)

B 55,358 0.186 (0.229) 16.28 (40.73)
H 25.000 0.564 51.96

Mean biomass of three visits.

- 591 -

with continuously flowing water as did trout farm B and
'appeared* to function effectively.

Phosphorus is the first limiting nutrient in the aquatic
environment, and in relation to eutrophication phosphate (P-PO^)
in the water is believed to be immediately available for the
growth of benthic algae (Alabaster, 1982). Bergheim et al.
(198M) observed that a considerable part of the TP in their study
was in this soluble reactive form (PO^-P). Soluble reactive
phosphorus was not measured in the present study, however on the
basis of Bergheim*s study it can be assumed that the level of
reactive phosphorus closely approximates the TP measured in this
study. Correlating the biomass of the trout farms with the total

daily phosphorus output of the farms indicated a highly

2
significant linear regression Cr = .79) with the following

equation.

y (phosphorus output) = 0.0146 x (farm biomass) - 2.04

This would suggest that it is possible to predict the phosphorus
loading from a trout farm on the basis of its biomass.
Obviously, it is desirable and perhaps essential that the
phosphorus output from the fish be kept to a minimum. As
previously mentioned the use of a retention and/or settling pond
may be beneficial in reducing phosphorus output but this remains
to be determined. Aside from reducing the biomass, which is not
really economical or realistic, probably the only other cost
effective way of reducing phosphorus output is to reduce the
phosphorus intake and excretion in the fish. This will be
discussed in a later section of this report.

- 592 -

In contrast to the relationship of trout farm biomass to
phosphorus output, surprisingly the relationship between trout
farm biomass and ammonia output did not appear to be closely
correlated (correlation coefficient r value of 0.22). Ammonia
is thought to be the major nitrogen excretory product of fish,
being approximately 90%, in fish such as trout (Brett and Groves,
1979). Therefore trout farm output would be expected to be
directly related to the biomass of the farm. The fact that it
does not in the present study is difficult to explain. Bergheim
et al . (1982, 198U) noted that of the total nitrogenous wastes
excreted by the fish farms, only 26% of the total nitrogen was in
the form of ammonia. These researchers did not really explain
this large difference other than to suggest that waste feed may
be accounting for some of the excess nitrogen. While there is
little doubt that some waste feed could be contributing to the
total nitrogen output, it is difficult to accept that it would
make up the large difference between total nitrogen and ammonia
output. However, it may suggest that other nitrogenous wastes
are being excreted by the fish reared on these trout farms.
Furthermore, the lack of correlation between biomass and ammonia
output on the trout farms in the present study may be due to the
increased excretion of other forms of nitrogenous wastes in the
trout. This obviously warrants further study.

Results

Experiment I

After 12 weeks on the test diets there was no significant

difference in the final body weight of trout reared on the test

- S97, -

TABLE 13

Final body weight, feed:gain ratio, mortalities and final
carcass proximate composition of the rainbow trout
after 12 weeks on the test diets (experiment 1)

Parameter Diet Number

1 2

79.9 78.4

1.2 1.1

0.8 0.8

51.8 51.7

40.4 40.1

8.8 7.9

1.3 1.3

Final Body Weight
(g/fish)

79.7

Feed: Gain

1.1

Mortalities (%)

Carcass Analysis
Protein
Lipid
Ash
Phosphorus

50.7

43.4

8.4

1.2

Results expressed on a dry weight basis.

- 594 -

diets (Table 13). Proximate composition analysis of the final
carcasses indicated no significant differences. Comparing the
different groups of trout, the phosphorus content of the final
carcasses was not significantly different.
Experiment II

After 16 weeks on the test diets there was a significant
decrease CP<.05) in the final body weight of trout reared on diet
7 as compared to trout reared on the control, diet U, (Table 1U).
There were no significant differences between the final body
weights of the remaining groups. Trout reared on diets 5-7
exhibited significantly lower (P<,05) feed:gain ratios than trout
reared on the control diet M. Mortalities {%) were not
significantly different (P>.05) between different groups of trout
and were less than 2% for any group.
Biochemical analyses

No significant differences were detected in the carcass
protein levels (Table m); however, the results were somewhat
variable. In contrast, carcass lipid levels were observed to be
significantly higher (P<.05) in trout reared on diets 5-7 than on
the control diet U, Similarly, carcass ash levels were also
significantly higher (P<.05) in trout reared on diets 5-7 as
compared to trout reared on the control diet U. There was no
significant difference (P>.05) in the carcass phosphorus level of
the various groups of trout (Table 14); however, the percentage
phosphorus retention was significantly higher in trout reared on
diets 5-7 as compared to the trout reared on the control diet M,
No significant differences were detected in the PER's (Table 15)

- 595 -

TABLE 14

Final body weight, feed: gain ratio, mortalities and final
carcass composition of the rainbow trout after
16 weeks on the test diets (experiment II)

Parameter

Diet
5

Number
6

Final Body Weight
(g/fish)

52,8^^

50.6^^

47.8^

50.9^^

Feed: Gain

1.2^

0.9**

0.9^

0.9*'

Mortalities (%)

Carcass Analysis
Protein
Lipid
Ash

48.4
41.1*
8,1*

44.6^
49.3^
10.3^

46.5.

47.2?

11.?

49.8^
45.9^

Phosphorus

1.3

1.2

1.3

Results expressed on a dry matter basis.

^Results with same letter superscript not significantly different
(P>.05).

- S96 -

or the hemoglobin and hematocrit levels of the different groups
of trout.

Discussion

On the basis of the growth parameters, final carcass
composition and phosphorus content (Table 13), supplemental
dicalcium phosphate is not required in low fish meal practical
trout diets. Since commercial trout diets normally contain in
excess of 25% fish meal in the diet formulation (Hilton and
Slinger, 1981), the endogenous phosphorus content of the diet
would already be in excess of the phosphorus requirement level of
0.7-0.8% of the diet (Ogino and Takeda, 1978). Furthermore, this
excess level of phosphorus would probably increase the total
phosphate and soluble P-PO^ content of the trout farm effluent.
Therefore, it is recommended that commercial trout diets
containing 25% or more fish meal should not be further
supplemented with dicalcium phosphate.

On the basis of the growth parameters in experiment II
(Table 1M), final carcass composition and phosphorus retention
(Table 1U), it appears entirely possible to increase phosphorus
retention and reduce phosphorus excretion in the trout by dietary
manipulation. By feeding diets containing higher levels of
protein and lipid than in the control diet 4 (Table 12), the
total feed consumption, as reflected by the significantly lower
feed:gain ratios (Table 14), of trout reared on diets 5-7, was
reduced. This resulted in a reduction in the amount of
phosphorus consumed by the fish from approximately 88 g/100 g
fish in trout reared on the control diet to approximately 60

- S97 -

g/100 g fish in trout reared on the test diets 5-7 over the 16
week experiment. The fact that both the control and test diets,
with the exception of diet 6, maintained essentially the same
growth parameters (Table lU), hemoglobin and hematocrit levels,
PER and carcass phosphorus content (Tables 14, 15) indicates that
both control and test diets supplied sufficient amounts of
available phosphorus. Ogino and Takeda (1978) determined that
the phosphorus requirement for rainbow trout was between 0,7 and
0.856 of the diet. However, this requirement level would be
dependent upon the feed intake of the trout and the available
energy content of the diet. Both the control and test diets in
this study (Table 12) contained in excess of the phosphorus
requirement as determined by Ogino and Takeda. However, the
reduction in feed intake of the trout on the high protein-high
lipid diet may result in a higher available phosphorus level
being required in the test diet. Further studies are required to
determine the phosphorus requirement of trout reared on these
high-protein, high-lipid diets. Nevertheless, the results of
this study demonstrate a principle which is of considerable
importance in the formulation of low pollution diets for rainbow
trout.

Three different test diet formulations (Table 12) were used
In this study to illustrate that a high-protein, high-lipid diet
could decrease feed and phosphorus consumption and decrease
phosphorus excretion, in the trout. However, the levels of
protein and lipid chosen in this study may not be the most
advantageous in terms of the growth, feed efficiency, cost
effectiveness and physiological response of the trout. The final

- S98 -

TABLE 15

Phosphorus retention^ and protein efficiency ratio (PER)
of the rainbow trout after 16 weeks on the test diets

(experiment 11)

Parameter

Diet Number
5

Phosphorus Retention (%) 18.3
PER 2.09

34.6
2.02

27.7
2,06

25.4
1.98

^Phosphorus retention - (Phosphorus retained * Phosphorus fed) x 100
2pER - Weight gain v weight of protein fed.

- 599 -

carcass composition of the trout reared on test diets 5-7
indicated a significantly higher level of carcass lipid than in
trout reared on the control diet (Table 14), Furthermore, trout
reared on diet 6 had a significantly lower final body weight than
did trout reared on the control diet. A more judicious selection
of protein and lipid levels, as well as feedstuffs used in the
formulation of the test diets, would appear possible for the
formulation of more cost effective, production efficient, low
pollution trout diets.

Phosphorus excretion in fish can occur by either the urinary
and/or fecal route. While dietary manipulation can obviously
increase phosphorus retention and decrease phosphorus excretion,
it cannot be determined from this study whether either or both of
these excretory routes were affected by this procedure. The
release of fecal phosphorus in the effluent water may be
dependent upon microbial action and therefore it seems possible
that urinary phosphorus may be the major source of waterborne
phosphorus in the trout farm effluent although that remains to be
determined. Future studies on low pollution trout diets should
determine whether both fecal and urinary phosphorus are affected
by dietary treatment, and which is the greater source of
waterborne phosphorus .

ACKNOWLEDGEMENTS

The authors wish to thank Ms. Debbie Conrad, Mr. Marty
Hodgson and the various graduate students and summer students
for their technical support during the collection and analysis of

- 600 -

this data; and the trout farmers surveyed in this study without
whose help and cooperation this project could not have been

completed.

i.L

- 601 -

REFERENCES

Alabaster, J.S. 1977. Biological Monitoring of Inland

Fisheries. FAOI Calliard Limited, Great Yarmouth, 226 pp.
Bergheim, A. and A.R. Selmer-Olsen. 1978. River pollution from

a large trout farm in Norway. Aquaculture 14: 267-270.
Bergheim, A., A. Sivertsen and A.R. Selmer-Olsen. 1982.

Estimated pollution loadings from Norwegian fish farms, I.

Investigations 1978-1979.
Bergheim, A., H. Hustveit, A. Kittelsen and A.R. Selmer-Olsen.

198M. Estimated pollution loadings from Norwegian fish

farms. IX. Investigations 1980-1981. Aquaculture 36: 157-

168.
Bligh, E.G. and W.G. Dyer. 1959. A rapid method of total lipid
extraction and purification. Can. J. Biochem. Physiol. 37:

911-917.
Brett, J.R. and T.D.D. Groves. 1979. Physiological energetics.

In 'Fish Physiology' Volume VIII, (eds. W.S. Hoar, D.J,

Randall and J.R. Brett). Academic Press, New York, pp. 279-

352.
Faure, A. 1977. Mise au point sur la pollution engendree par

les pisicultures. Pisiculture 13: 33-35.
Milton, J.W. and S.J. Slinger. 1981. The nutrition and feeding

of rainbow trout. Can. Spec. Pub. Fish. Aquat. Sci. 55:

15p.
Hinshaw, R.N. 1973. Pollution as a result of fish cultural

activities. Ecological Research Series, U.S. Environmental
Protection Agency, Washington, D.C. 20 460, 53 PP.

- 602 -
Horwitz, W. 1980. Offical Methods of Analysis of the Assoc, of

Anal. Chem. Thirteenth Edition, AOAC, Washington, D.C.

200^4.
Ogino, C. and H. Takeda. 1978. Requirements of rainbow trout

for dietary calcium and phosphorus. Bull. Jap. Soc. Sc i ,

Fish. 44: 1015-1018.
Standard Methods. 1980. For the examination of water and

wastewater. 15th edition. American Public Health

Association , Washington , D.C. , 1 134 pp.
Steel, R.G.D. and J.H. Torrie. 1980. Principles and Procedures

of Statistics: A Biometrical Approach. 2nd ed . pp. 172-

288. McGraw-Hill Book Co., N.Y.
Tervert, D.J. 1981. The impact of fish farming on water quality

Wat. Pollut. Cont. 1981: 571-581.

- 603 -

DESIGN OF GPOUNDWATEP MONITORING PPOGRAMS
FOP WASTE LANDFILL SITES

Richard P. Schwarz Philip H. Byer

Graduate Student Associate Professor

Department of Civil Enaineerinq , University of Toronto

- 604 -

ABSTRACT

The increased concern over qroundwater ouality monitorinq at
waste landfills has resulted in many advances beinq made in
samplinq techniques, contaminant transport modellinq and
leqislation for qroundwater protection. However, little work has
been done to systematically desiqn monitorinq proqrams by
incorporatlnq into desiqn procedures samplinq costs, data
variability, environmental damage costs and probabilities of
contaminat ion .

The research to be reported here demonstrates how a
hierarchy of models of increasinq complexity for monitorinq
proqram desiqn can be built up from basic hydroqeoloq ical ,
statistical, econom ic and optimization concepts. The monitorinq
proqram desiqn is for detection-type monitorinq concerned with
observinq siqnificant chanqes in water quality, and where
remedial action is initiated upon detection. The tradeoff
between periodic monitorinq cost and the increased remedial
action costs resultinq from undetected plume orowth is expressed
mathematically. The effect on the total cost of monitorinq and
remedial action by decision variables, such as the number of
monitorinq wells, the frequency of monitorinq and the
siqnificance level of the statistical test used for impact
assessment, is invest iqa ted.

- 605 -

INTRODUCTION

One of the basic objectives for monitoring groundwater
quality at waste landfill sites is to detect the occurrence of
contamination. Such monitoring typically requires the periodic
analysis of groundwater samples taken from monitoring wells which
are downgradient and adjacent to the landfill. The measured
chemical concentrations can then be compared with the background
water quality data using a statistical test to determine if there
has been a significant impact on the water quality. A finding of
statistical significance, when verified by a retesting procedure,
then leads to some form of remedial plume corrective action-
This is the basic philosophy of the USEPA-RCRA (United States
Environmental Protection Agency-Resource Conservation and
Recovery Act) groundwater regulations (Schwarz and Byer , 1984),
Figure 1 shows the relationship of the background and monitoring
wells to the landfill.

The research reported here develops mathematical models to
facilitate the planning of these detection monitoring programs.
The following characteristic features of these programs are
considered in the models:

(1) Periodic monitoring costs, which are expensive and may last
for decades, can be quantified in terms of the time period
between samples, the number of wells being sampled, the discount
rate and the time until contamination arrives.

(2) Since the monitoring program is being designed for uncertain
eventualities, it is necessary to use probability theory. A
discrete probability measure is assigned to the likel ihood of
whether contamination will occur during some time horizon. A
continuous probability distribution describes when this
contamination will occur, conditional on contamination occur ing
at all. The sampling and hypothesis testing procedures which are
used to decide whether contamination is present utilize the
concepts of independence and normality of the water quality data.

(3) Consistent with the USEPA-RCRA groundwater rules, and in
order to achieve a clear objective for the mathematical
optimization procedure, remedial action is assumed to be
initiated immediately upon detection of contamination.
Furthermore, there is no advance warning that contamination will
soon arrive at the monitoring wells.

(4) The cost of performing plume remedial corrective action,
which depends on the hydrogeologic and economic conditions at a
particular site, is expressed as a function of the time permitted
for undetected plume growth.

(5) when comparing different time streams of monitoring costs
and plume correction costs, a discount rate is used to quantify
the time value of money.

This paper develops the above concepts into a system of decision
models. Initially the monitoring program design is formulated
for the following assumptions; (a) no variability in the sampling

- 606 -

FIGURE 1

LANDFILL-MONITORING SYSTEM

groundwater
flow

future plume of
contamination

WASTE

background
wells

monitoring
wells

PLAN VIEW

background
well

\^ pi ume

» / monitoring well

bedrock

ELEVATION

- 607 -

data, (b) a discount rate of zero, (c) contamination is certain
to eventually occur, and (d) not allowing for the possibility of
systematically revising the monitoring program as time goes by.
These restrictive simpl if ying assumptions are then progressively
relaxed to make the model more realistic.

MONITORING PROGRAM DESIGN OBJECTIVE

The overall objective used to establish a landfill site and
initiate a monitoring program is to maximize the discounted net
benefits, where the benefits result from society's willingness to
pay to dispose waste at the site (minus the site operation
expenditures ) and the costs arise from periodic monitoring and
plume remedial action (if contamination occurs). Since the times
of arrival and detection of groundwater contamination, if it
occurs at all, are uncertain, the resulting discounted net
bene fits are governed by a probability distribution. Using
analytical expressions, the discounted expected net benefits are
calculated and used to compare alternative monitoring program
designs. In spite of the advances being made in analyzing
stochastic cash flows {Barnes et. al., 1978 and Zinn et. al.
1977), it is not considered useful here to attempt to derive and
use more information about the probability distribution of the
discounted net benefits.

In order to include the effect of the landfill benefits on
the monitoring program design, for the general model development,
it is assumed that the site is closed on detection of
contamination and that the time horizon of when contamination is
no longer probable coincides with the anticipated closure date of
the site. The resulting equations are easily modified to analyze
the monitor ing and remedial action programs when the landfill
benefits are not considered commensurable or when they occur
independently.

The time plots shown in Figure 2 help to illustrate the
basic structure of the assumed decision process. The notation
used in this paper is defined in an appendix. The time axis is
discretized into monitoring intervals (A t) and into years. There
is an initial cost of establishing the monitoring program (cw), a
periodic monitoring cost (cm), and an annual landfill site
benefit ( v). At some unknown time in the future ( tc) groundwater
contamination may occur. Depending on the statistical test used
to detect contamination and its significance probability i^) , the
number of monitoring wells (nm) and the monitoring period (At),
there will be a stochastic time required to detect contamination
(td) with its attendant cost of remedial action (cr). The
anticipated life of the landfill and the upper limit on the
probability of when contamination might occur is shown as th.

On comparing examples A and B, it is seen that the sampling
interval Is increased which results in lower monitoring costs.
However, the time required to detect contamination and the

- 608 -

FIGURE 2

STOCHASTIC REALIZATIONS OF EXAMPLE MONITORING POLICIES

Example A (nm=2)

At At
cw cm cm

cm cm

' A '

cm , cm
tc

cm+cr
td

Example B (nm=2)

cm
tc

cm+cr
td

Example C (nm=4)

ent

cm ^
tc

cm+cr
td

Example D (nm=4)

At ■ A
cm ^
tc

cm+cr
td

- 609

resulting cost of remedial action is expected to be greater. A
tradeoff between monitoring cost savings and increased remedial
action cost is evident. The difference between examples B and C
is that the number of monitoring wells has been increased from
nm=2 to nm=4. Consequently, the expected time to detect
contamination will be less although the increased periodic cost
of sampling 2 additional wells may not be worth the earlier
detection. The difference between examples C and D is that due
to chance, contamination has occurred sooner. There have been
fewer years to realize the annual landfill benefit and the
present worth of the remedial action cost is larger. The
modelling procedure deals with the above described costs and
stochastic variables in a mathematically rigorous fashion.

COST CONCEPTS

cr = a + b(td) (1)

where or = cost of remedial action, $

a = fixed cost {independent of plume size), $

b = variable cost of plume growth, $/year

td = detection time after contamination begins, years

Cost estimates for various alternative plume remedial strategies
have been developed (Geraghty and Miller, 1982), and thus (1)
represents the cost function of the least costly alternative.
The physical extent of the plume {the major variable cost factor)
is embodied in cr by relating the groundwater velocity and plume
dispersion to the time available for undetected plume growth.

A fundamental assumption is made when using (1). The rate
and shape of plume growth is considered independent of when the
plume starts growing, which is governed by a probability
distribution. Furthermore, it is assumed that the rate and shape
of the plume growth can be predicted sufficiently accurately that
the cost of remedial action may be estimated. For the case where
the model incorporates sample uncertainty, it is assumed, in
addition, that the concentration of contamination is known, or at
least a value can be specified for design purposes. The
assumption being made here of independence between the plume
characteristics and the time when the plume arrives cannot be
sufficiently emphasized since it models independently what
contaminant transport modelling typically models conjointly.
This assumption of independence was introduced by Sobotka (1983)
and appears to be conceptually consistent with the growing body

- 610 -

of literature on the economics of groundwater pollution (see
Raucher, 1983 and Kavanaugh and Wolcott, 1982). As possible
extensions, it would be mathematically feasible to build in a
functional dependence between contaminant transport model
parameters and the time of contamination arrival as well as to
assign probability distributions to the various plume
characteristics such as width, velocity and chemical
concentration .

For remedial action occurring at a given year t in the
future, the present worth using continuous discounting is,

-rt

P = cr e (2 }

where P = present worth of plume remedial action cost
occur ing t years in the future , $
t = future year

i = discount rate , % per annum
r = equivalent continuous discount rate
= In (l+i)

The monitoring costs are assumed to consist of initial
installation costs and periodic monitoring costs as follows
(Duvel Jr., 1982) ,

cwv = (nm + nb) ciw + cmo + cmv (3)

where cwv = total well installation cost, $

nm = number of monitoring wells

nb = number of background wells

ciw = cost of installing a well, $/well

cmo = fixed mobilization cost of installing wells, $

and

cmv = ( nm + nb ) csw + cms { 4 )

where cmv = monitoring cost for uncertain sampling, $/period
csw = unit cost of sampling from a well , $/well
cms = fixed cost of mobilizing for sampling , $

The modifier v on cwv and cmv indicates that the formulas
are used for model cases which incorporate sample variability. A
modifier c is also used (cmc) to indicate sampling with
certainty. This cost is discussed in the section on sampling
under uncertainty and is a retesting cost set equal to a simple
multiple of cmv. The simpler model cases (which do not consider
sample uncertainty) utilize the costs cw and cm , which do not
make the monitoring cost explicitly dependent on the number of
wells- In addition, these equations do not explicitly include
consideration of the number of chemicals to be analyzed for and
the number of samples to be taken from each well since the
statistical hypothesis testing procedure to be used is not
sufficiently refined to need this type of cost detail. The costs

- 611 -

FIGURE 3

UNIFORM PROBABILITY DENSITY FUNCTION FOR TC

f (tc)
Tc

1
th

PI / '

f (tc)= 1 ; < tc < th
Tc th

PI = tl/th = PCtc <. tl) •

■ tl th

Time of Contamination (tc), years

FIGURE 5
PLUME REMEDIAL ACTION COST FUNCTION FOR SAMPLING UNDER CERTAINTY

Remedial

Plume

Cost(cr)

a+bAt

a .

2At 3At 4At 'th-2At th-lAt

Time of Contamination (tc), yed,r^

- 612 -

are referred to as cw and cin for qeneral discussion or where the
context is clear.

For a period of t years with sanplinq every At years, the
total undiscounted monitorina cost is,

C(total) = cw + (cm t)/At (5)

With continuous discountinq, the present worth counterpart of (5)

is,

e - 1

C( total) = cw + cm \ rAt M ^^ j ^^*

PPOBARTLTTY CONCEPTS

Three separate forms of probability are used in the model.
A probability (pc) is specified in advance ("a priori") as the
probabil ity that contamination will eventually occur. This
probability can be considered as the deqree of belief in the
existence of a state (ie. eventual con tarn ination). This state
probability is defined as follows,

P( e = eu) = pu (7 )

p( e = ec) = pc (8 )

pu = 1 - pc (^)

where oc = the state that contamination will occur before
time th
Ou = the state that contamination does not occur
before time th
pu,pc = the probabilities of the respective states

Tt is relatively easy to modify this deqree of be lief (also
commonly referred to as a subjective probability) as time passes
without the occurrence of contamination.

The second recourse to probability theory also occurs in
advance and provides the probability of when contamination will
occur, qiven that it will eventually occur (state ec). For this
paper the uniform continuous distribution is used since it
results in simple mathematical relationships and has an intuitive
appeal when iudqement is needed to estimate when contamination
miqht occur. This distribution is shown in Fiqure 3. The third
probability concept previously mentioned deals with the
hypothesis testinq for the difference of the mean water quality
between up and downqradient wells and is discussed in a later
section on samplinq under uncertainty.

Fiqure 4 shows the structure of the time framework of the
probability updatinq procedure which can be used to modify the

FIGURE 4

DECISION TREE FOR PROGRAM UPDATIN;

t=0 Atl
(Now)

Atl

. ENB(tc,Atl)

Contamination Before tl]

■Nr

tl At2
(Reassessment Point)

lt2

ENB(itl,it2)

(No Contamination)

^ ENB(tc,Atl,At2)

(Contamination After tl)

(^ Chance Node

Cvj Decision Node

ENB Expected Net Benefits

tc Contamination Arrival Time

Ptc Probability of tc

N 1.

- 614 -

monitoring program at a future time tl. Consider the situation
at time t=tl when contamination has not been observed and where
the sampling does not involve uncertainty about the presence of
contamination ( in the sense that statistical errors com pi icate
the statement that contamination has not been observed). Since
the eventual occurrence of contamination was initially uncerta in
(pc '-. 1 ) , it is reasonable that this subjective probabil ity might
be modified if contamination has not been observed up to a
particular time. This reassessment is referred to as posterior
analysis in Bayesian terminology (Benjamin and Cornell, 1970).
The posterior probability for the eventual occurrence of
contamination (still within the time horizon th) given that
contamination has not yet occurred is.

pc' = P(x 1 ec)pc (10)

p(x| ec)pc + p(xi eu)pu

where pc'= posterior probability of eventual contamination
X = the elapsed time that contamination has not
been observed

Note that since th has not been adjusted, pc serves as the
probability that contamination will occur during the time segment
t2, given that it has not occurred during tl.

For the uniform probabil ity distribution, the density
function is given by,

f (tc) = 1/th ; Oltc^th (11)

where tc = the time when contamination occurs, years
th = the upper limit on the time of occurrence
of contamination, years

The expected value of when contamination will occur is,

E(tc) = th/2 (12)

Given that contamination will eventually occur (within time th),
the probability that it will occur before time tl is,

P(tc < tl) = tl/th (13)

The unconditional probabilities, assessed at t=0, that
contamination will occur during the time segments tl and t2 are,

P(tc < tl i t=0) = (tl/th) pc (14)

P(tc > tl I t = 0) = (t2/th) pc (15)

- 615 -

Note that (14) oivfis pel of Finure 4, The conditional
probability distributions for when contain inat ion will occur
within these time seqnents are,

f (tc|tc<tl) = (l/th)/{tl/th) = 1/tl (16)

f (tc|tc>tl) = l/t2 (17)

The conditional nxpochations arp,

R(tc |tc<tl) = tl/2 (IR)

K(tc|tc>tl) = tl + t2/? (19)

Solvinq for the posterior probabil ity of (10) qives,

pc' = (1 - tl/th)pc (20)

(1 - tl/th)pc + (l)pu

where pc = pc2 of Fiqure 4 .

STOCHASTIC CASH FLOWS

Tt is necessary to calculate the present worth of time
streams of monitorinq costs, landfill benefits and a future plume
remedial action cost over an uncertain time duration, Tt is not
correct to use roqular discount inq equations with the expected
time duration. Fquations have been derived which qive the
expected present worth for various cost trends and probability
distributions (Zinn et.al. 1977). The followinq formulas pertain
to discountinq cash flows where the time duration, tc, is
uniformly distributed. For a compoundinq period of 1 year and
usinq the continuous discountinq approximation, the expected
present worth of a series of R dollars per year (eq. v or cm
( r/(exp( r At )-l ) ) ) over tc years is,

f-r th
1 + e - 1 1 (21)

r th

The expected present worth for a lump sum of S dollars occurrinq
at the random time tc is.

(22)

It is important to realize the S and R may also be the expected
values of random future costs.

- 616 -

EXPECTED PLUME REMEDIAL ACTION COST

For sampling under certainty, there is a relatively simple
relationship between the expected plume remedial action cost, cr,
the probability distribution for tc. At, r and the plume cost
parameters a and b. For sampling under uncertainty, the ability
to detect contamination will also depend on the power of the
hypothesis testing procedure, this is discussed in a later
sect ion .

The interrelationship between the monitoring interval ( At) ,
the plume remedial cost (cr) and the time of contamination (tc)
is shown in Figure 5. The effect of the monitoring interval, for
certain sampling, is to limit the plume cost to the range a <, cr
< a + bAt. The expected plume remedial action cost is obtained
by integrating Figure 5 with the probability distribution for tc,
(11).

At 2 at

/ (a + b(At-tc))f (tc)dtc + /

E(cr) = / (a + b(At-tc))f (tc)dtc + ( (a + b(2At-tc))f (tc)dtc

Tc
At

(a +b(th-tc))f (tc)dtc (23)

Xc
th-At

This integral is solved, for the general case , by isolating the
general integral ,

i At

Ei = / (a + b(iAt-tc))f ( tc ) dtc (24)

Tc
(i-l)At

and using analytical summation eguations to perform the
summation,

i=(th/At)

E{cr) = > Ei (25)

i = l

For the uniform distribution with zero discount rate,

E{cr) - a + b At/2 (26)

This is intuitively appealing since given a uniform distribution

- 617 -

Cor Lc, the probability distribution for the start
contamination within a sampling period must also be uniform.

f (tp) = 1/At
Tp

; <_. tp SAt (27)

where tp = the time of contamination arrival within At

For the analysis which divides the time horizon into two
time segments (tl and t2), and for which At will be chosen
separately (Atl and At2), the loss function of Figure 5 may be
integrated according to the conditional probability distributions
(16) and (17). For the uniform distribution with zero discount
rate ,

E(crl tc < tl)
E;(cr| tc > tl )

a + b A tl/2
a + b A t2/2

(28)
(29)

Although the integration procedure yields the obvious results of
(26) to (29), the results for other probability distributions are
not as obvious. For the case of zero discount rate, E(cr) of
(26), (28) and (29) are modified to include the annual land fill
benefit by subtracting v from b. Since v is treated as a
continuously discounted cash flow while b is only discounted once
contamination is detected, subtraction of v from b is not
permitted for non-zero discount rates.

For the uniform probability distribution with an annual
discount rate (r), and a single monitoring interval (At), E(cr)
is obtained by discounting a , b, cm and v to a uniformly
distributed original time within the monitoring interval. For
the discounting, a, b, and cm are treated as future values and v
is treated as a continuous cash flow.

E(cr| t=tc)= I -

r(At-tp) \ / \ -r(A t-tp)

e ^1 + ((a+cm)+b(A t-tp)

r(At-tp)

.dtp (30)

-rAt -rAt -rAt

= (a + cm) (1 - e ) + b (1 - e - rAte )

rAt 2

r At

-rAt
- v (r At + e - 1)

2
r At

- 61!

R(cr lt=hc) is treated as a randomly occurrino cost and is
discounted to t=0 usinq (22),

-r th

E(cr|t = n) = E(cr|t = tc) f 1 - e \ (31)

r th

RXPFCTED NFT RFNEFTTS - MONITORING UNDEP CERTAINTY

The expected net benefits are calculated by addinq the
expected costs of plume remedial action, E(cr), the periodic
costs and benefits ,cm and v, incurred either until
contamination occurs, or for the lifetime of the site, and the
initial costs of install ina the monitoring system, cw. The
ex pec tod net' benefits are then used to provide a qauqe for
select i no the best moni tor inq program,

CASK (a) : Consider the simplest case: a uniform probabil ity
distribution for tc, certain contamination (pc=l) , zero discount
rate (r=0), certain samplinq, and no reassessment of the
monitorinq policy after tl years. The expected time until
contamination is qiven by (12) , and the expected plume remedial
action cost by (26). The expected net benefits (FNB) are then,

FNB = -cw + Kv - cm )\ th - ( a + cm + (b-y)At. ) (32)

\ tt I 2 2

To find the optimal monitorinq interval, the derivative of FNB
with respect to At is set to zero , qivinq.

At = X cm th (33)

V h-v

For example, if the monitorinq costs are $S000 per samplinq
period, contamination is expected sometime in the next 25 years,
the marqinal cost of plume remedial action is $2 00,000 per year
of undetected plume growth and the annual landfill benefit is
$100,000 then the optimal monitorinq interval is every 1.12
years. Should v exceed b, then (33) indicates that there is no
incentive to perform monitorinq. Retting v=0 and making the
costs neqative in ENB transforms (33) to that of minimizinq the
total costs of monitorinq and plume remedial action, reqardless
of the landfill benefit. The square root form of (33) and the

- 619 -

"sawtooth" profile of Figure 5 suggest an analogy with the
"economic lot- size model" used in inventory theory (Hillier and
Lieberman, 1980).

CA.SE( b) : Consider the assumptions of case(a) but relax the
assumption of certain contamination (ie. pc<l).

ENB = -cw + pc

(v- _cm ) _th, - ( a + cm + {b-v)A_t) )
A t 2 2

+ (1-pc) ( V th - cm th )

(34)

Maximizing ENB with respect to At gives,
At =

y cm th /_2.- l\
b-v ^ pc ;

(35)

Continuing the numerical example started in case( a) , if pc= 0,5,
then the optimal monitoring interval becomes A t = 1.94 years.

CASE( c) : It may be <Josirable to reassess the monitoring interval
at some prespecified time in the future (at year tl). The
assumptions are uniform tc, pc<l, certain sampling, discount rate
r=0, and tl+t2=th. The decision tree (Fig. 4) indicates that the
optimal monitoring intervals Atl and A t2 are decided upon at
times t = and t = tl respectively. The ENB at t = are,

ENB = -cw + pel ENB(tC<tl) + pul pc2 ENB(tC>tl)

+ pul pu2 ENB (no contamination) (36 )

where pel = ( tl pc)/th

pul pe2 = (t2 pc)/th
pul pu2 = 1-pe

Using the conditional expectations for the time of arrival of

contamination
and (29), ENB

ENB = -cw +

(18) and (19)
is g iven as.

and the conditional plume costs (28)

pc tl ( -a -cm - (b-v) A tl
tH \ 2

+ V tl

- cm

+ pc

+ (1-pc) ( V th - cm tl - cm t2
\ Atl At2

(37)

Differentiation of ENB with respect to Atl and At2 and setting
to zero gives ,

- 620 -

Atl =

cm
(b-v)

2
At2 =

(" '{h") j

-tl +(_2
(b-v) I Vpc

fe--)

- 11 th

(38)

(39)

Comparinq (38) with (35) it is seen that if t2=0, that is no
reassessment of At is beinq considered, then the two equations
are identical, Continuinq the numerical example from case(b) , if
it is anticipated that the men i tori no proqram should be modified
at the midpoint of the time horizon for contamination
( tl = t2 = l 2.5) , then Atl = 2.09 years and At2 = 1.77 years. Thus if
after 12.5 years contamination has not been observed, then it is
optimal to monitor more frequently. On comparinq (3R) with (39),
it is seen the At 2 is always less than Atl. This conclusion, when
compared with the formulations for other distributions for tc ,
such as the downward trianqular probabil ity distribution, is seen
to be due to the uniform distribution beinq used here, which
results in pc2, (20), increasinq over pel. It is slqniflcant to
point out that maximizinq PNB at t=tl for At2 qives the same
result as (39).

CASK(d) : The final case to be presented for the certain sampl inq
assumption utilizes the discount rate. Tc is taken as uniform,
only a sinqle At is considered, and pc is allowed to vary. Iisinq
(6) for the discountinq of annual costs and benefits when no
contamination occurs, (21) for the discountinq of annual costs
and benefits when contamination does occur and (31) for the
expected plume costs and benefits after the arrival of
contamination qives FNB as follows.

r th

ENP = -cw + (1-pc) (v - cm

[T^lll-^

- 1

r th

-r th

)

-r th f -rAt -rAt

■*■ (]-e ) |v(rAt+e -1) - b(l-e (l + rAt))l

r t- h A t

+ ( _v - cm

r rAt

e -1

(40)

- 621 -

Al thouqh it: does not appear feasible to differentiate (40) with
respect to At, plotting FNP vs. At or usinq a numerical search
procedure, should readily qive the optimal At. Tn contrast to
cases (a), (b) and (c), the fixed plume remedial action cost, a,
will affect the selection of At since, as a result of the
discountinq, it is economically advantaqeous to postpone
incurr inq this cost.

RXPECTRD NET BENEFITS - MONITOT?ING WITH UNCERTAINTY

The rosul ts qi ven in the last two sect ions are derived under
the assumption of certain sampMnq. This implies theri^ is no
chance of makinq false conclusions due to chance variation of
the water quality data. This variability is a result of samplinq
and laboratory analysis procedures as well as the heteroqeneous
nature of the qroundwater quality. Two types of error are of
concern: (1) concludinq that contamination is present when in
fact it is not (in statistical terminoloqy, a false positive or
Type 1 error), and (2) concludinq that contamination is not
present when in fact it is (a false neqatlve or Type 2 error).
Although the presence of variability complicates the analysis of
water quality data, common statistical tests can be used to help
interpret the data. For the monitorinq problem, a typical
concern is with establishinq the existence of a siqnificant
difference between the qroundwater quality above and below the
landfill (Rovers and McBean, 1981). Since the impact of data
variability on the cost and performance of the monitorina pronram
may be considerable, it is desirable to incorporate this aspect
into the program desiqn.

For the case of uncertain samplinq, the model defines two
states of nature correspondinq to: (1) the time when the
qroundwater is not contaminated (60) and (2) when the qroundwater
is contaminated (el). These states are dist inquished from those
describinq whether contamination will eventually occur (ie.
eu= the state that qroundwater contamination will never occurand
state ec = the state that qroundwater will eventually occur.
Statesai and ec are qoverned by pc as shown in (7), (8) and (9)
while 80 and 9 1 are qoverned by the probability d i str ibu t ionf or
tc.

A statistical test is made at the time interval At and a
decision HO is made that the qroundwater quality is eO or a

decision HI is made that the qroundwater quality is 91. Table 1
describes the relationship between the states of nature, the
statistical decisions and the probabilities of makinq false
conclusions .

- 622 -

TABLE 1 : TRUTH TABLE FOR HYPOTHESIS TEST

State of
Nature

Test Indication

fiO Hi
(Uncontaminated) { Contaminated)

(Uncontaminated)

( Contaminated )

1 - a a

( False Postive)

6 1 - P
(False Negative)

a= The probability of inferring that there is contamination when
none exists .

p= The probability of inferring that there is no contamination
when it is present.

Due to the variability of the water quality data and as a
consequence of the statistical hypothesis testing procedure, two
additional cost concepts are required. As a result of the
possibility of false positive errors, it is necessary to require
a retesting procedure to verify that contamination is not
present. For the purpose of this model, the retesting procedure
is assumed to produce certain knowledge, but at a cost greatly
exceeding the sampling cost which produces uncertain information.
This cost is referred to as cmc ( the cost of monitoring with
certainty) and may involve extensive retesting at the monitoring
wells, the use of other sources of information such as
geophysical surveys, greater numbers of chemicals being analyzed
for or perhaps more sophisticated water quality analysis
techniques being used. The probability that this cost will arise
at each sampling period, when there is no contamination, is n ,
and thus an expected cost of acme must be added to the routine
monitoring cost cmv of (4).

The second additional cost concept involves the additional
cost of plume remedial action which occurs when false negative
errors are made. The probability that a false negative occurs,
when contamination is present, at each sampling time is given by
B . Since these errors may, by chance, occur consecutively,
Bernoulli sampling theory (Benjamin and Cornell, 1970), is used
to calculate the probability of the possible sequences of these
errors. A relationship is derived between the sampling interval ,
the probability of not detecting contamination and the increased
plume remedial action cost.

The basic solution procedure is essentially the same as with

- 623 -

cort. ain sampling, Tho costs and benefits are calculated
separately for the time periods before and after the arrival
contamination. Although the discounting considers when the
contamination might occur, it is assumed that the actual future
cost of plume remedial action is independent of when the
contamination occurs. However, this future cost is not
independent of how long it takes to detect the contamination
after it occurs .

Table 2 shows a possible decision sequence over time. At
time 5At, contamination has arrived at one or several of the
monitoring wells, as shown by the change from GO to 9 1. Row H
shows a possible sequence of hypothesis test indications, with
the probabilities of their occurrence being given underneath.
For the hypothesis test of the model, the test indications are
made independently in time. This implies that a subjective
probability on the presence of contamination is not being updated
and that trend analysis of time series data is not being
performed. Also shown in Table 2 are the costs associated with
making errors in the hypothesis testing. At time 2At , a false
positive occurred and the cost cmc was required to confirm that
contamination was not present. At times 5At and 6At,
contamination was present but had not been detected, causing an
increase in the remedial action cost (Acr). At time 7At the
contamination is detected and cmc (or perhaps a fraction of cmc)
is needed to validate the result. At this point remedial action
would take place.

A relatively simple statistical hypothesis test is utilized
here to help illustrate the methodology. The hypothesis test
analyzes the difference of means of two normal populations having
a common known variance, for a s ingle contaminant. The
statistical testing may be extended to multiple contaminants by
using Hotelling's T^ (Davis, 1973 and Bickel and Doksum, 1977),
and to cases where the variance is estimated from the data by
using the non-central student's t distribution.

TABLE 2: EXAMPLE SEQUENTIAL DECISION PROCESS

Time

lAt

2At

3At

4 At

5At

6At

7At

State

Probabil ity

1- a

1-a

1- a

1-a

1-P

Error Cost

cmc

Acr

cmc

For the model development of uncertain monitoring, it is
assumed that the contamination will arrive as a step increase
over background and that the magnitude of the step is known in

- 624 -

advance. Normally distributed data are assumed,

Cbd ^ N(yu,
Cmd '^^ N(u u,
Cmd '\> N(Mc ,

a2 )
a2 )

t < tc
t > tc

(41)
(42)
(43)

whore Pu = mean value of uncontaminated groundwater data

Mc = mean value of contaminated groundwater data

a^ = variance of groundwater data

Cbd = background concentration data

Cmd = monitoring ( downgradient ) concentration data

The difference of the estimated means is therefore normally

distributed as follows,

(Cmd - Cbd) -x- N( 0, a^/nm + aVnb)
(cind - Cbd) % N( 6 , a^/nm + aVnb)

6 = p c - y u

; t < tc
; t > tc

(44)

(45)
(46)

Eq.'s (44) and (45) assume that the data are independent. For a
right-sided hypothesis test, the two hypotheses are (Guttman et.
al., 1971),

pm - M b
pm - P b

=
>

(47)
(48)

where

ub = mean of the background data
\im = mean of the monitoring data

For notational consistency, ub is
represents the same data mean as yu.
rejected if.

Cmd - Cbd
o^ /nm + oVnb

> z

introduced although it
The null hypothesis is

(49)

Note

„w^^ that for this test, the data averages Cmd andCbd are
estimated from the data, but that a^ is assumed known. Assuming
a known variance may not be unrealistic given that data may be
collected over long periods of time. The probabilities of the
two types of errors are ,

Cmd - Cbd

o-ynm + "■ /nb

m -ub =

(50)

- 625 -

where $(z) is the area under the standard normal probability
curve from - » to z. Figure 6 shows the relationship between the
standard normal distribution and the error probabilities. It is
seen that a is specified independently and, in particular, does
not depend on the number of wells, whereas B depends on many
tactors as given in (52). Figure 7 shows the general shape ot a
power tu notion. The figure shows that the probability ot
detection increases as the Level ot contamination increases.
Although it is desirable to increase the power, such as by making
nm and nb larger, this additional expense must be justified, as
is done in the subseguent model development.

Although it presents no conceptual problem to make (4 3) time
dependent to allow for a diffuse contaminant arrival front, the
consequence of the step front arrival is that 3 remains constant
and thus the number of sampling periods required to detect
contamination is distributed as a geometric random variable,
G(l-3), (Benjamin and Cornell, 1970).

P (N=n) =

n-1

(1-B)

n = 1,2,3,

(53)

where N = number of sampling periods until detection
B = probata il ity of not detecting contamination
(1-6 ) = probability of detecting contamination

The time required to detect contamination is related to the
monitoring interval At, to N of (53), and is also related to the
probability distribution of tc within the sampling interval At-
Since tc is taken as uniform, it is seen that the occurrence of
contamination is also uniform within At, (27). The time required
to detect contamination is made up of two independent random
variables ,

td = (At- tp) + (N-1 ) At

54)

where td = time required to detect contamination

tp = time of contamination arrival within At

- 626 -
FIGURE 6

STANDARD NORMAL DISTRIBUTION AND TESTING ERRORS

FIGURE 7
THE POWER FUNCTION

1.0

POWER
(1-6)

1-B = l-*(z^ -

a2 + o-

6 = yC - yu

- 627 -

'Che cn^t oF remedial action nnd tho monit.orino costs and landfill
bonofits which arise after contamination occurs are qoverned by
(54). Whore discountinq is used, it is done in two staqes, first
to td = (ie. t = tc), and then to t = n usinq (22). Two model cases
are developed for uncertain monitorinq, with and without
discountinq .

CASE(e) ; This model case is for uniform tc,samplinq with
uncertainty, no revision of At, pc variable and zero discount
rate. The expected costs that are incurred after the arrival of
contamination are ,

E(cr) = (a + cmc) + (1-B) ((b-v) At/2 + cmv)

+ (1-B) ((b-v) 3At/2 + 2cinv)

2
+ B (1-B) ((b-v) 5At/2 + 3cmv) + ... (S5)

The basis for calculatino F(cr) is that a fixed cost a and
certainty monitorinq cost cmc are incurred, and the annual costs
b and benefits v are incurred accordinq to the probability of
successive occurrences of false neoatives precedinq the
successful detection. Usinq the qeometric sumTnation equations,

2 3

a + ap + ap + ap + ... = a/(l-p) ; |p| < 1 (S6)

P + 2p(l-p) + 3p(l-p) + ... = 1/p ; <, p< 1 (57)

to evaluate (55) qives ,

F(cr) = (a + cmc) + (b-v) At n/2 + B/(l - 3 )) +cin v/( 1 -B ) (58)

Separatinq ENP into the net benefits with and without the
eventual occurrence of contamination, and includinq the
monitorinq costs and benefits until the arrival of contamination
o ives ,

mc) th ( Sq )

vt h- ( cm v+ acme) t h
2 2At J

Maximizinq ENB with respect to At qives,

A^- = J cmv -f acmc)th(2 - pc ) (60)

pc(b-v) 1+B
1-B

- 628 -

Note thc\t sottinn a = n= fl in (60) qives (35). Since RNB is also
affected Uy a , nm and nh (indirectly throuqhg , Fq.(52) ), they
can also be considered as decision variables for maximizina ENP,

CAPR(f ) : This case relaxes the assumption of zero discount rate
used Tn case(e). Tn order to discount the plume remedial action
costs, monitor! no costs and land fill benefits over the time
required for detection (td), the following integrations are used.

-r( At-t) / r( At-t)

a+cmc+b( At-t)| e - v /e

-']

r( At-t)/ At

-r(2At-t) / r(2At-t)

a + cmc+b( 2At-t)| e - v f e ) I dt^

r(2At-t) 1 I At
re

-3

/-r(2At-t)
cmv e d^

(61)

Discounting (61) to t=0 using (22), and including the monitoring
costs and benefits up to the occurrence of contamination gives.

+ pc

v^ - ( cmv+acmc)
r rAt

(e -1)

-rth

1 + e 2^

rth

- 629 -

-rth

+ 1 - e

rth

-rAt
-( (1-B ) (a + cinc)+CTnv) {1-e )

-rAt
rAt(l- Be )

-rAt

rAt{l-Be )

-rAt

(l-B)be

-rAt
krd-pe )

B-1

-rAt

l-pe

rAt
e ^

rAt

(62)

The decision variables of interest in Tnaxiinizinq (62) are a ,
At,mTi and nb. Since B may be obtained from a hiahly accurate
analytical approximation to the error function, (62) is
essentially analytically solved.

ILLUSTRATIVE EXAMPLES

This section demonstrates the use of the expected net
benefits models derived in the preceding sections. The annual
landfill benefit v is taken as zero so that the objective is to
minimize the expected total cost which consists of monitoring and
plume remedial action. Hypothetical data are used to illustrate
the form of the results and to demonstrate the types of tradeoffs
which are characteristic of groundwater monitoring programs.
Example 1 is concerned with monitoring under certainty and
Example 2 with monitoring under uncertainty.

EXAMPLE 1: Model cases (a), (b) and (d) are used to solve
five combinations of th, pc and i, but which otherwise use common
data as shown in Figure 8. Since the main design parameter of
concern under the certainty assumption is the monitoring
interval, its affect on the expected total cost is graphed.

With the exception of combination C, a monitoring interval
may be found such that a minimum total cost is found. The effect
of increasing the discount rate in going from A to C is to
increase the monitoring interval to such an extent that there is
no incentive to monitor. Due to the time value of money,
postponing remedial action costs is economically preferred.

When comparing A and D, the effect of increasing the
probability of eventual contamination is seen to increase the
expected total cost and to decrease the optimal monitoring

- 630 -

FIGURE 8: RESULTS OF EXAMPLE 1

1.3

1.2

■H

-H

1.1

0.0

o
o

o.a

0.7

0.6

0.5

liJ
□L

0.4
0.3

0.2

0.1

MONITORtNCS INTERVAL (YEARS)
B o C A D

COMBINATIONS

years

100

COMMON DATA

a = $500,000, b = $100,000/year,
cm = $6,250/moni toring interval

%/annum

5
10

cw = $28,250,

- 631 -

FIGURE 9: RESULTS OF EXAMPLE 2

c
o

•H

■H

0.9

O.B -

O 0.7 -

O 0.6

S 0.5 -

0.4

D A

-B-

-I 1 1 1

4 6

NUMBER OF MONITORING WELLS
o C a

COMBINATIONS

csw

mgVl^

$/wen

500

750

%/annum

5
5

COMMON DATA

a =- $500,000. b = $200.000/year. ciw - SSOOO/well . cmc=3cmv.
cms - $3000/mom'toring interval, pc = 0.75, th = 25years, pc
yu = 25 nig/1

cmo ^ $1000.
30 mg/1 ,

- 632 -

interval. The increased likelihood of contamination has resulted
in more effort being expended to rapidly detect it. The result
of increasing the time horizon of contamination occurrence is
seen, when comparing A and E, to increase the optimal At.
Although the probability of the eventual occurrence of
contamination is the same, since it is spread out over a longer
time horizon, the same intensity of monitoring effort is not
oconnmically warranted. Although the results of Fxamplo 1 depend
on thp particular cost and probability parameters used, and thus
extracting general conclusions is not possible, the optimal
monitoring interval is seen to adjust to input data in an
intuitively satisfying way.

EXAMPLE 2: Model cases (e) and (f) are used to solve five
combinations of a , a^, csw and i, as shown in Figure 9, The
expected total cost is shown as a function of the number of
monitoring wells (nm), which, in this example, is also equal to
the number of background wells ( ie. nb=nm) . In showing the total
cost for a given nm, the least cost monitoring interval (in
years) was taken from the set {0.25, 0.33, 0.5, 0.75, 1, 1.5,
2, 2.5, 3, 4, 5, 6, 7, 8, 9, 10). In the example, it is common
for the monitoring interval to increase as nm increases with the
typical range being from 0.33 to 1.5 years.

Figure 9 shows that in this example the expected total cost
is not very sensitive to the number of monitoring wells when the
least cost monitoring interval is being selected. Each plot
shows that a minimum total cost is achieved with nm usually in
the range of 3 to 5 wells, although for B the minimum occurs at 8
wells. The range of highest to lowest total cost for the five
plots , as nm varies, is in the range of $20,000 present worth.
A greater range in the expected total costs could have been
achieved by adjusting the cost and probability parameters to
cause more drastic tradeoffs. Table 3 shows how the minimum cost
changes as nm and At vary. It is noted that varying nm, for a
fixed At, may cause excessively large expenditures.

TABLE 3: EXPECTED TOTAL COST FOR COMBINATION C, EXAMPLE 2

{$ MILLION)

At (years)

0.25

1.012

1.080

1.152

0.33

0.913

0.960

1.012

0.50

0.833

0.853

0.880

0.75

0.822

0,817

0.824

1.00

0.853

0.829

0.821

1.50

0.957

0.902

0.870

2.00

1.082

0.999

0.946

- 633 -

The effect of changing the significance probability a is seen
when A and B are compared. Since amay be considered a decision
variable, for the particular data being used here, it is
preffirablo to use a =0,05 rather than 0.01. The tradeoff between
thoso two values of ais that a smaller value of "decreases the
power of tho statistical test, resulting in greater remedial
action costs, but also leads to fewer false positive errors which
require additional monitoring expenditure. In this example, this
cost was taken as being 3 times the normal monitoring cost (cmc =
3cmv) .

Comparing combinations C and D shows the effect of
increasing the variance of the water quality data. The total
cost increases as a result of the greater difficulty in detecting
the contamination. Although the figure shows that the optimal
number of wells is less for D than for C {4 vs. 5), the optimal
monitoring interval was less for D (0.5 years vs. 0,75 years).

Combination E increases the per well sampling cost to
$750/well from $500/well used in C. This results in the optimum
for E having fewer monitoring wells (4 vs. 5 for C). The optimal
monitoring interval is 0,7 5 years for both C and E,

CONCLUSIONS

"I h(j (level opmont and application of a pi ann ing model for
designing groundwater quality monitoring programs is presented.
The model demonstrates the interaction between sampling and
environmental damage costs, the probability of the occurrence of
contamination and monitoring decision parameters such as the
number of monitoring wells, the monitoring frequency and the
desired statistical reliability. Among the relationships which
have been studied, the following conclusions can be made;

(1) Increasing the probability of the eventual occurrence of
contamination and decreasing the time horizon of when it will
occur leads to greater monitoring expenditure .

(2) Based solely on econom ic considerations, large benefits
resulting from landfill operation can override the need for plume
remedial corrective action, if it is a question of closing the
landfill on detection of contamination.

(3) Given a set of cost values and contamination probabilities,
the monitoring interval and the number of monitoring wells may be
adjusted to achieve an overall minimum cost for monitoring and
plume remedial action .

(4) The practice of discounting future costs leads to lower
monitoring expenditures, and in certain cases, may make it
uneconomical to monitor at all,

(5) Increasing the cost of plume remedial action, especially the
marginal cost of undetected plume growth, will lead to greater
monitoring effort. More frequent monitoring is also caused by

- 634 -

lower sampling costs and by highly variable water quality data.
(6) ihe specification of a desired statistical reliability is
seen to depend on the tradeoff between the cost of retesting the
monitoring wells and the cost of undetected plume growth.

The overall performance of the model is seen to yield
results which are intuitively satisfying. For the probability
distribution assumed here for illustrative purposes, relatively
simple analytical equations are obtained which can be used to
improve the design of monitoring programs. Further work is being
done to incorporate other probability distributions, multiple
contaminants and more frequent reassessment of the program
design ,

REFERENCES

1. Barnes, J. Wesley, C. Dale Zinn and Barry S. Eldred, " A
Methodology for Obtaining the Probability Density Function
of the Present Worth of Probabilistic Cash Flow Profiles",
AIIE Transactions, Vol, 10, No. 3, 226-236, 1978.

2. Benjamin, Jack R. and C. Allin Cornell, " Probability,
Statistics and Decision for Civil Engineers", McGraw-
Hill, New York, 1970.

3. Bickel, P.J., and K.A. Doksum, "Mathematical Statistics:
Basic Ideas and Selected Topics", Holden-Day, San Francisco,
1977.

4. Davis, John C, "Statistics and Data Analysis in Geology",
John Wiley and Sons, New York, 1973.

5. Duvel Jr., William A., "Practical Interpretation of
Groundwater Monitoring Results", Proc. National Conference on
Management of Uncontrolled Hazardous Waste Sites, Washington
D.C., Hazardous Materials Control Research Institute,

Nov. 29-Dec. 1, 1982.

6. Geraghty and Miller, Inc., "Cost Estimates for Containment of
Plumes of Contaminated Groundwater", prepared for U.S.
Environmental Protection Agency, Annapolis, Maryland, 198 2.

7. Guttman, Irwin, S.S. Wilks and J. Stuart Hunter , "Introductory
Engineering Statistics", John Wiley and Sons, New York, 1971.

8. Hiilier, Frederick S. and Gerald J. Lieberman, "Introduction
to Operations Research", 3rd. Edition, Holden-Day,

San Francisco, 1980.

9. Kavanaugh, M., and R.M. Wolcott, "Economically Efficient
Strategies for Preserving Groundwater Quality", prepared for
U.S. Environmental Protection Agency, Public Interest
Economics , Washington D.C.,1982.

10. Raucher, Robert L. , "A Conceptual Framework for Measuring the
Benefits of Groundwater Protection", Water Resources
Research, Vol.19, No . 2, 320-326 , April 1983.

11. Rovers, F.A. and E.A. McBean, "Significance Testing for
Impact Evaluation", Groundwater Monitoring Review, Vol.1,
No. 2, 39-43, Summer 1981.

- 635 -

12. SfTliwarz, R.B. and P.H. Byer , "Summary and Discussion ot U.S.
EPA-RCRA Groundwater Rules", Solid and Hazardous Waste
Management Series, WM 84-09, Department of Civil Engineering ,
University of Toronto, January 1984.

13. Sobotka and Company Inc . , "The Benefits ot Avoiding Ground-
Water Contamination at Two Sites in the Biscayne Aquifer" ,
submitted to Office of Policy Analysis , U.S. EPA,
Washington D.C. , November 1983.

14. Zinn, CO., W.G. Lesso and B. Motazed, "A Probabilistic
Approach to Risk Analysis in Capital Investment Projects",
The Engineering Economist, 22, 4, 239-260, 1977

APPENDIX-NOTATION

a = fixed cost ot plume remedial action, $

a = significance probability of statistical hypothesis test

b = variable cost of plume remedial action, $/year

6 = probability of making a false negative statistical error

cbd = background concentration measurement , mg/1

ciw = unit cost of installing a well , S/well

cm = periodic monitoring cost , $/sample period

cmc = cost of monitoring with certain sampling

cmv = cost of monitoring with uncertain sampling
cmd = monitoring (downgradient) concentration measurement, mg/1
cmo = fixed cost of mobilizing for installing wells, $
cms = fixed cost ot mobilizing for sampling wells, 5
cr = total cost ot plume remedial action , $
csw = unit cost ot sampling from a well, $/well
cw = total initial cost of install ing well system, $

cwv = well installation cost tor uncertain sampling, $
6 = actual difference in the chemical concentration between

contaminated and uncontaminated groundwater( uc -Mu=5),mg/1
ENB = expected net benefits (present worth) of the landfill

site-monitoring program, $
HO = hypothesis that groundwater is not contaminated
Hi = hypothesis that groundwater is contaminated
i = discount rate , % / annum

N = number of sampling periods until detection of contamination
nb = number of background monitoring wells
nm = number of monitoring wells

pc = a priori probability of eventual contamination
pc' = posterior probability of eventual contamination given that

it has not occurred by year tl
pel = probabil ity that contamination will occur during tl
pc2 = probability that contamination will occur during t2 given

given that it has not occurred during t 1
pu ^ a priori probability of no eventual contamination
pul = probabil ity that contamination will not occur during 1 1
pu2 = probability that contamination will not occur during t2

given that it has not occurred during tl
r = equivalent continuous discount rate, % / annum
o = standard deviation of the groundwater quality data, mg/1
tc = time of contamination, years
td = time required to detect contamination, years

- 6:^6

At = monitoring interval, years

Atl = monitoring interval for time segment tl
At2 = monitoring interval for time segment t2
th = time horizon for landfill life and probable contamination,
years

tl = first time segment of th
t2 = second time segment of th (tl+t2 = th)

= state that groundwater is not contaminated

1 = state that groundwater is contaminated

oc = state that groundwater will eventually be contaminated
nu = state that groundwater will not eventually be

contaminated
tp = time of contamination arrival within the monitoring

interval, years
yb = mean concentration of the background wells, mg/1
yc = mean concentration of the contaminated groundwater, mg/1
ym = mean concentration of the monitoring wells, mg/1
yu = mean concentration of the uncontaminated groundwater, mg/1
V = annual benefit of having the landfill, $/year
X = elapsed time that contamination has not occurred, years
z = standard normal random variable whose probability of being

exceeded is a

ACKNOWLEDGEMENTS

This work was undertaken as part of a research grant from
the Ontario Ministry of the Environment (Provincial Lottery
Fund). Special thanks are extended to Irmi Pawlowski of the
Ministry for her constructive comments and interest in the study.

TD Proceedings ; technology

172.5 transfer conference no, 5

.057 76017

1984

part 1

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