Note: Descriptions are shown in the official language in which they were submitted.
~~~94~~
SY8T8M TO REDOCB BEDIMENT T08ICITY
FIELD OF THE INVENTION
The present invention relates to a method and apparatus for
the treatment of sediment and more particularly to a chemical
treatment method and apparatus to effect such treatment.
HACICaROUND OF T8g INVENTION
Generally speaking, most of the biodegradation studies and
treatments are done in reactors (Jafvert and Rogers 1991). The
largest reactor is a Dutch system where a mobile aerator moves
along the treatment pond to maintain oxic conditions (Van Veen
and Annokkee 1991). Other large treatments include landfarming
(Van Dillen 1991) and the excavation and construction of a hill
of contaminated soil with drainage tiles, nutrient dosing and
sprinkler systems (Litchfield et al. 1992). These systems all
require that the sediment be dredged and processed in sites which
can involve the potential problems of security, cost, and public
acceptance.
Some soil reclamation using biodegradation has occurred
without excavation. The success of these treatments has varied
greatly (Lee et al. 1988). Some contaminants cannot be readily
biodegraded. Some treatments may not have provided the right
environmental conditions for the microbes to completely
biodegrade the organic wastes. Most of the failed bioremediation
treatments in groundwater are related to either poor access to
the wastes or to the blockage of the aquifer by the enhanced
growth of microbes. The latter two problems do not apply to lake
sediment on a macroscale: however, all sites can contain some
polynuclear aromatic hydrocarbons (PAHs) locked in microsites
which are therefore not biodegradable (Van Dillen 1991).
Presumably these refractory PAHs would not be toxic. In similar
studies it was found that coal dust contains PAHs but relative
to coal tar and creosote, it is biologically inert (Alden and
Butt 1987).
Bioremediation of groundwater and soils is a growing
industry. As long as the treatment is preceded by an analysis
of the treatability and toxicity of the site, it is a promising
2
remedial option. Some toxins cannot be treated by
biodegradation, but the cost of assessment is justifiable in that
a detailed preassessment is substantially less expensive than
excavation and chemical or physical treatment.
Bubbling the water column with oxygen has been proposed as
a method of oxygenating the sediment (Murphy 1990). Some lake
aeration treatments in Germany successfully oxygenated sediments
but treatment with pure oxygen of some lakes in Switzerland did
not oxygenate sediments (Gachter 1987). The engineering
techniques are not completely developed, the treatment time may
be long, and recovery difficult to predict accurately.
Further prior methods for treating sediment of sludge
include the method of decontaminating wastewater sludge taught
by Nicholson in U.S. Patent No. 4,781,842. Reference is made to
the use of lime, cement kiln dust and lime kiln dust for treating
biological sludge. The goal of the treatment. is to fertilize
agricultural land. The reference does not address the problem
of treating sludge associated with a water body for the purpose
of enhancing natural microbiological degradation.
U.S. Patent No. 5,008,020, discloses the use of a metal
carbonate and a metal bicarbonate for solidifying waste material
into a granular particulate form. The patentees do not discuss
the merits of oxidation biodegradation inhibitors in order to
effect more efficient biodegradation.
Conover, in U.S. Patent No. 5,039,427, teaches a method for
removing suspended solids and to precipitate and inactivate
phosphorus in lake water by adding aluminum hydroxide sulphate.
The reference does not discuss oxidation of sulphide to sulphate
to reduce inhibition of natural microbial biodegradation.
U.S. Patent No. 4, 877, 524, teaches an apparatus for treating
bodies of water for correcting chemical, biological or other
imbalances. The reference primarily relates to dispensing a
3
treatment agent within a water body, the dispensing rate being
proportional to the boat speed. Both aluminum sulphate and
sodium aluminate are taught as treatment material among others.
There is no provision in the apparatus for inj ecting or otherwise
contacting the sediment with biochemical oxidant and accordingly,
no contemplation for enhancing the natural microbial degradation
of the sediment and more particularly the toxins of the sediment.
Shuck et al. in U.S. Patent No. 4,268,398, teaches a method
for rendering a sludge deposit in a waste water treatment
facility into a pumpable mixture for relocation. This reference
does not discuss an in-situ procedure for natural biodegradation
of the sludge.
Further generally related references include U.S. Patent No.
5,055,204 and Canadian Patent Application Nos. 2,007,455 and
2,016,310
In view of the prior art, there exists a need for a sludge
treatment process capable of being performed in-situ without
stirring up the sediment into the water column and which permits
a relatively large area to be treated quickly.
80Z~MARY OF THE INVENTION
One aspect of the present invention is to provide an
improved method for treating sludge and an improved apparatus for
effecting the method.
In accordance with another aspect of one embodiment of the
present invention, there is provided, a method of effecting
natural microbial biodegradation of polynuclear aromatic
hydrocarbons and petroleum hydrocarbons in sediment containing
microbes and polynuclear aromatic hydrocarbons and petroleum
hydrocarbons, comprising the steps of: providing a biochemical
oxidant for detoxifying a microbial toxin produced during
microbial biodegradation of the polynuclear aromatic hydrocarbons
and petroleum hydrocarbons: contacting the sediment with the
oxidant to detoxify the toxin; and effecting microbial
biodegradation of the polynuclear aromatic hydrocarbons and
petroleum hydrocarbons.
The process of microbial decay results in the production of
hydrogen sulphide. The presence of this compound generally
impedes the biodegradable of polynuclear aromatic hydrocarbons
and petroleum hydrocarbons. The addition of a biochemical
oxidant; e.g. a bivalent alkaline earth chloride or nitrate or
monovalent nitrate have been found useful.
In accordance with another aspect of the present invention
there is provided a method of effecting natural microbial
biodegradation of polynuclear aromatic hydrocarbons and petroleum
hydrocarbons in sediment containing the polynuclear aromatic
hydrocarbons and petroleum hydrocarbons and microbes, comprising
the steps of: providing a biochemical oxidant for oxidizing a
biodegradation inhibitor produced during microbial biodegradation
of the polynuclear aromatic hydrocarbons and petroleum
hydrocarbons; contacting the sediment with the oxidant;
oxidizing the inhibitor to a non-inhibiting form: and effecting
microbial biodegradation of the polynuclear aromatic hydrocarbons
and petroleum hydrocarbons.
By oxidation of the inhibitor, substantial success was
realized in the reduction of toxins in the sediment. The
successful results were compounded by the fact that the sediment
was contacted at a plurality of locations.
The overall process is continuous such that the treatment
of the sludge can be effected quickly over a large area.
Nutrients may be added with oxidant concurrently or
conterminously. Further, the addition of emulsifier surfactants,
or other suitable treatment aids may be employed. Depending on
the specific site variables, a pretreatment may be desirable in
order to increase the efficiency of the procedure.
5
Apparatus suitable for facilitating maximum exposure of the
oxidant to the sediment includes a dispensing arrangement
therefore for dispensing the oxidant into the sediment in at
least two locations simultaneously.
Thus, according to a further aspect of the present invention
there is provided an apparatus suitable for treating sediment in
a water body, the sediment containing chemical pollutants,
comprising: a first dispensing means adapted for dispensing a
treatment compound into contact with the sediment to be treated
at a first level; a second dispensing means level adapted for
dispensing the treatment compound within the sediment at a second
level; each of the dispensing means having means for connection
with a supply of the treatment compound; and support means for
supporting the first dispensing means and the second dispensing
means.
A further aspect of the invention is to provide an apparatus
suitable for treating sediment in a water body, the sediment
containing chemical pollutants, comprising in combination: a
first dispensing arrangement for dispensing a treatment compound
into the sediment at a first level; a second dispensing
arrangement for dispensing the treatment compound into the
sediment at a second level, each the dispensing arrangement
having means for connection with a supply of the treatment
compound; a support frame for connecting the first dispensing
arrangement and the second dispensing arrangement; and a carrier
vessel for the vessel including means for connection and
manipulation of the support frame.
Having thus generally described the invention, reference
will now be made to the accompanying drawings illustrating
preferred embodiments.
$RIEF D88CRIPTIODT OF T8B DRAWINC~B
Figure 1 is a perspective view of one embodiment of the
apparatus of the present invention;
~~~942~
Figure 2 is an enlarged perspective view of one embodiment
of the sediment treatment apparatus;
Figure 3 is a section along line 3-3 of Figure 2;
Figure 4 is an enlarged view of the mounting and fluid
distribution for the sediment treatment apparatus;
Figure 4A is a plan view of the sediment treatment apparatus
support;
use;
Figure 4B is a schematic illustration of the apparatus in
Figure 5 is a series of graphs illustrating the
concentrations of various contaminants;
Figure 6 is a schematic illustration of the test area;
Figure 7 is a graphic illustration of the redox potential
of St. Marys River sediment cores;
Figure 8 is a histogram illustrating the effect of ferric
chloride on sediment hydrogen sulphide for the St. Marys River
treatment;
Figure 9 is a histogram illustrating the ATP-TOX results
from Sault Ste. Marie for the first ferric chloride sediment
injection;
Figure 10 is a series of histograms illustrating the
sediment toxicity of photobacterium phosphoreum bioassay for data
gathered at St. Marys River;
Figure 11 is a histogram illustrating the a is ~g~a
bioassay results illustrating the average percent survival;
~~~J~2
Figure 12 is a graph illustrating the percent mortality for
the DMSO/methanol sediment extract:
Figure 13 is a histogram illustrating the average percent
survival for Daphnia magna bioassay results for the Hamilton
Harbour Stelco Hotspot site #1;
Figure 14 is a histogram illustrating the average percent
survival for Daphnia magna bioassay results for the Hamilton
Harbour Stelco Hotspot site #2;
Figure 15 is a histogram illustrating the average percent
survival for Daphnia magna bioassay results for the Hamilton
Harbour Stelco Hotspot site #3;
Figure 16 is a histogram illustrating the results of Figures
13, 14 and 15;
Figure 17 is an illustration of the toxicity severity of
Hamilton Harbour;
Figure 18 is an illustration of the toxicity of Hamilton
Harbour sediments for photobacterium;
Figure 19 is an illustration of the Hamilton Harbour
sediment injection site;
Figure 20 illustrates data generated in denitrification
experiments for bottle incubations;
Figure 21 illustrates data generated in denitrification
experiments for 250 ml bottle incubations;
Figure 22 illustrates data generated in denitrification
experiments for 2 L bottle incubations;
g
Figure 23 illustrates data generated in denitrification
experiments for 250 ml bottle incubations;
Figure 24 illustrates data generated in denitrification
experiments for 2 L bottle incubations;
Figure 25 illustrates the biodegradation of the sixteen
priority pollutant polynuclear aromatic hydrocarbons;
Figure 26 illustrates the denitrification data comparing
Hamilton Harbour Deep Basin, Stelco Hotspot and St. Marys River;
Figure 27 illustrates the effect of the nitrate treatment
on the Photobacterium;
Figure 28 illustrates headspace GC/MS analysis for a variety
of organic compounds for a control and for data gathered after
nitrate treatment; and
Figure 29 illustrates the biodegradation of volatile toxins
in the Dofasco boatslip before and after treatment.
DETAILED DEBCRIPTION OF THE PREFERRE EMBODIMENTS
Referring initially to the apparatus aspect of the present
invention, Figure 1 illustrates a first embodiment of the present
invention, in perspective, as situated on a carrier vessel.
Generally, the apparatus includes rotatable lifting
apparatus 10 having a main support platform 12 mounted to the
carrier vessel 14 in a known manner. Load bearing masts 16, 18
are connected to the platform 12. A winch cable system 20
includes winches 22 extends over pulleys 24 provided on the load
bearing masts 16 and 18. One of the winch cables 20 includes a
connecting member 26 suitable for connection with the sediment
treatment apparatus, generally denoted by numeral 30. A support
system is provided for supporting and assisting in the
~~~~~23
positioning of the treatment apparatus 30, and will be discussed
hereinafter.
In one embodiment of the treatment apparatus, best
illustrated in Figure 2 and 3, an elongate hollow spray bar 32
is included having opposed ends 34 (Figure 1) and 36. The spray
bar 32 includes a plurality of nozzles 38 distributed along the
length of bar 32 in a spaced and aligned relation. The nozzles
38 are inclined downwardly relative to a horizontal plane. A
second set of nozzles 40 are distributed along the length of
spray bar 32 in spaced and aligned relation radially spaced from
nozzles 38. Both nozzles 38 and 40 are in fluid communication
with spray bar 32.
As is illustrated in the example, the sediment treatment
apparatus 30 is composed of a plurality of connected and similar
units and accordingly the description will be limited to one such
unit.
Spaced above spray bar 32 there is provided a mounting
member 42 comprising a metal tube having opposed ends 44 (Figure
1) and 46. A plurality of such members 42 are connected in end-
to-end relation by suitable fasteners as illustrated. Each
remote or terminal end as well as the end connection between
mounting members 42 includes a spacer member 48, which includes
a spacer member 48, which not only spaces the members 42 from the
spray bar 32, but additionally serves to impart support to the
spray bar 32. Connection of spacer members 48 to member 42 and
bar 32 is achieved by suitable fasteners, welding etc.
Each mounting member 42 includes a plurality of arcuate
fingers 50 each connected thereto by bolts 52. The fingers 50
are arranged in longitudinally aligned and spaced relation. Each
finger 50 subscribes to a generally sinusoidal configuration and
each comprises a rigid metal suitably bent into the indicated
shape. The shape of each finger 50 permits resilient
10
flexibility. A free end 54 of each finger 50 is laterally and
vertically spaced from the spray bar 32.
Each finger 50 includes, spaced from free end 54 thereof,
a nozzle mounting 56 for mounting a nozzle 58. A conduit 60
extends between and connects nozzle 58 with nozzle 40 such that
fluid communication is established.
Treatment fluid is distributed to each spray bar 32 by a
distribution conduit 62 connected to each spray bar 32 inwardly
of the end thereof by a swivel type connector 64 well known to
those skilled. This is generally illustrated in Figure 4. Each
conduit 62 terminates for fluid connection with a main feeder
conduit 66 (Figure 1). Conduit 66 is connected to a fluid
treatment supply drum 68 centrally located on the carrier vessel
to act as ballast. A pump (not shown) may be positioned
intermediate of supply drum 68 and feeder conduit 66 or the fluid
may be distributed by negative pressure.
Treatment fluid travelling through each spray bar 32 will
be dispensed through nozzles 38 as well as through nozzles 40,
all conduits being in fluid communication.
At least the spray bars 32 adjacent the terminal sections
thereof further include frame mounts 70 for releasably coupling
the treatment apparatus 30 to a frame 72.
Figure 4 illustrates an enlarged view of the attachment
of the end of frame member 74 to spray bar 32 (removed for
clarity) as is generally illustrated in Figure 1. Fastener 76
links a flange 78 on member 74 to frame mount 70. The frame 72
may include a plurality of members 74 which converge, and the
member thereof will vary depending on the size of the treatment
apparatus. The frame 72 permits easy manipulation of the
apparatus from a submerged position to a storage position, the
latter position being illustrated. Frame 72 includes a
11
connection site 80 for connection with connecting member 26 on
winch cable 20.
Figure 4A illustrates the hangar assembly 90 for
positioning the assembly 30 into the sediment at the desired
position.
The assembly 90 includes a main load bearing member 92
which terminates at a horizontally disposed bracket 94 for
to connection with spray bar 32 (not shown). Connection to spray
bar 32 may be by any suitable means, e.g. clamps, bolts etc.
Bracket 94 and member 92 are further reinforced by
braces 96 extending therebetween.
In order to permit treatment of sediment in a variety
of situations where the depth requirement varies, load bearing
member 92 may include a plurality of telescopic sections of tube
98 or may be extended by progressive manual connection of further
lengths of tubing sections 98.
Load bearing member 92 may be pivotally connected to
the load bearing masts 16 and 18, described herein previously,
or may be connected directly to platform 12 for easy
manipulations of member 92. The specific mounting position of
member 92 will depend upon the specific parameters in the
treatment area, e.g. water depth etc.
Generally, the tubing sections 98 may comprise rigid
3o aluminum material or other suitable corrosion resistant
materials. This material provision is additionally applicable
to the overall assembly 90 and frame 72.
In operation, the treatment apparatus 30 is moved from the
storage position shown in Figure 1 to the.use position shown in
Figure 4B where the apparatus is submerged below the surface of
the water, W, to contact within the sediment bed, S.
12
In this position, both the spray bars 32 and the fingers 50
contact the sediment, S, as illustrated. The fingers 50 permit
deeper penetration of the nozzles 58 and more specifically the
treatment material dispensed therethrough, into contact with the
sediment. The nozzles 38 dispense the treatment material in a
second position spaced from that of the treatment supplied by
nozzles 58. This two-point injection system has a dramatic
effect on the sediment detoxification as well be evinced by the
data discussed hereinafter.
The treatment apparatus 30 is dragged along the sediment bed
as the carrier vessel travels the area to be treated. The
arcuate fingers 50 are particularly advantageous for the sediment
treatment since the same are resilient and basically unaffected
by irregular bed topography, small debris etc. When encountered,
the fingers 50 simply flex and return to a normal disposition as
the apparatus continues to be advanced along the sediment bed.
Recovery of the apparatus from the sediment may be achieved with
the winch system described previously.
In preferred form, the overall length of the apparatus is
eight (8) metres with the spacing between nozzles 38 and 58 being
between 10 and about 20 centimetres. Such spacing permits
uniform dispersion within the sediment as opposed to localized
areas of treatment.
The rate at which the treatment fluid is injected into the
sediment may be timed with the carrier vessel speed i.e. a higher
vessel speed will require a higher rate of injection of fluid
into the sediment. In an alternate embodiment, the treatment
apparatus 30 and the ancillary equipment (winch, loud bearing
masts etc.) as well as the submerging procedure may be effected
by robotics controlled from the shore or at a point distant from
the treatment area. This arrangement would reduce the exposure
of human workers to the hazardous sediment material and presence
around the heavy equipment.
13 . ~~~94~~
The use of monitoring means e.g. sonar equipment, cameras
ultrasonic equipment etc. are all envisioned for use with the
apparatus in order to monitor gross sediment topography
irregularities or obstacles with which the apparatus 30 cannot
contend.
Still further, the connection points between spray bars 32
and between mounting members 42 may be hinged to permit folding
of the apparatus 30. In addition, the apparatus may be
telescopic.
The examples illustrated teach a two-point treatment
injection system and it will be clearly understood that a
multiplicity of injection points may be provided simply by, for
example, the addition of a further series of fingers having a
greater length than the previous series.
Having thus described the apparatus, reference will now be
made to the experimental procedures and generated data.
a$NBRAL
$~~~PLB 1 FBRRIC CHhORIDB INJBCTIO~if
Earlier laboratory trials with Hamilton Harbour sediments
indicated that the addition of iron reduced toxicity to
Photobacterium phosDhoreum, Daphnia manna, Sa
lmo crairdneri,
Pimenhales promelas, and Hexa enia limbata. The seasonal change
in sediment toxicity also seemed related to a change in redox,
albeit the relationship was not firmly established. Also there
was a correlation between the toxicity of the sediments to
Daphnia ma_qna and the chemical oxygen demand of the sediments.
The most appropriate hypothesis to explain these observations is
that much of the acute toxicity of the sediments of Hamilton
Harbour was caused by reduced chemicals, probably hydrogen
sulphide.
Hydrogen sulphide is very toxic. The LCSO for various
species are: Assellus 1.07 mg/L, Cran~,g~onvx 0.84 mg/L, Gammarus
..a~.,,,
~~9~~
14
0.059 mg/L, Baetis 0.020 mg/L, Ephemera 0.361 mg/L, and Hexagenia
0.111 mg/L. Chronic analysis indicates that no-effect levels are
about 10% of the LCSO (Oseid and Smith 1974). Although hydrogen
sulphide toxicity is well known, few studies report it in
sediments. US EPA (1986) describe hydrogen sulphide as
"ephemeral" which indicates that infrequent sampling would not
measure hydrogen sulphide. Another limitation is that some
professional laboratories use procedures too insensitive to
detect toxic concentrations of hydrogen sulphide.
Studies with 35S-radiolabels have measured the geochemical
reactions of sediment sulphur well. For example, Nedwell (1980)
determined sulphate reduction in summer is 50-100 times faster
than in winter. With intensive monitoring, Ripl (1986) observed
large seasonal changes in sulphate; if only a few percent of the
seasonal change in sulphate were converted to hydrogen sulphide,
the sediments were very toxic. This simplistic assumption may
underscore the ecological importance of seasonal changes in
hydrogen sulphide toxicity. Ingvorsen and Jorgensen (1982)
' observed a 20-fold seasonal change in sulphate reduction, but it
was matched by seasonal changes in hydrogen sulphide flux from
the sediments of 103-104 fold! At low rates of hydrogen sulphide
production, it was mostly adsorbed to particles, but at high
rates of hydrogen sulphide production, the binding sites were
saturated and the pulsed release of hydrogen sulphide would have
killed many benthic organisms.
Similar sulphur geochemical cycling and sediment toxicity
must occur in the hotspots of the Great Lakes but it is not
documented. Decades ago, all steel mills discharged large
quantities of sulphur from several sources such as sulphur balls
from the coking process or spent sulphuric acid in pickling
liquor. The areas in Hamilton Harbour and the St. Marys River
where these wastes would settle are anoxic and hydrogen sulphide
should form.
15
Inadequate documentation of the sulphur biochemistry could
lead to misinterpretation. Hydrogen sulphide has a half life of
about 19~19 h (Table 16 in Zehnder and Zinder 1980). Thus it is
easy to lose the toxicity by sample handling, such as is commonly
done when bubbling invertebrate bioassays with air. The reported
absence of Hexa enia from the St. Marys River has been explained
by correlation analysis to reflect oil and grease toxicity (St.
Marys RAP 1992), but it could easily be caused by hydrogen
sulphide toxicity.
Samples were collected on several trips to St. Marys River
and several trips to Hamilton Harbour with Ponar (Sault) or
Shipek (Hamilton) grab samplers, TechOps corers, and sediment
traps. The sediment traps were deployed a metre about the
sediments to determine if the sediment injection equipment
resuspended sediments. In the St. Marys River, four traps were
set at one upstream site and four traps were set at one site
downstream of the treatment area.
All St. Marys River samples were stored in a cold room at
the Great Lakes Forestry Research Centre, and processed for
shipping on ice, i.e., cores were extruded there at 1 cm
intervals. The Eh and pH of samples were recorded at the site.
All Hamilton Harbour samples were brought back to the institute
within hours of sample collection. All samples were chilled and
processed quickly, i. e. , all bioassays were processed within days
of sample collection. Sediment samples were subsampled; half was
freeze dried for metal analysis and half was frozen and retained
for organic analysis.
The ATP-TOX method of Xu and Dutka (1987) was used on 10%
DSMO 10% methanol elutriates. Equal volumes of sediment and DMSO
were mixed together and shaken vigorously by hand for 2 minutes.
The homogenized slurry was then centrifuged for 20 minutes at
10000 rpm. This system uses the measurement of ATP as indication
of microbial growth. If when compared to a control, a sample
16
inhibits ATP production (i.e., growth), a toxic effect is
assumed.
Daphnia magna bioassays were done on aqueous elutriates.
Within two weeks of collection all samples were extracted with
equal volumes of distilled water on an end-over-end shaker for
16 h. After extraction, sediment extracts were centrifuged for
20 min at 1000 g. Elutriates were centrifuged, not filtered.
Filtration can remove colloidal material that would not settle
from disrupted sediment and that may contain toxic metallic or
organic contaminants. Ten Daphnia less than 24 h old were
introduced to 25 mL of test medium and placed in a 25°C incubator
for 48 h. A 16 h light and 8 h dark photoperiod was used. Prior
to all experiments, pH and dissolved oxygen were measured and if
the oxygen concentration was less than 8 mg/L, the sediment
extracts were bubbled with purified air for 16 h. If more than
10% of the control Daphnia died within 48 h, the experiment was
repeated.
Photobacterium bioassays were run on whole sediments
(Brouwer et al. 1990). Dilutions for LCSO analysis were done
with clean sediments from an organic rich sediment from a marsh
near Long Point, Lake Erie.
BEDIMENTB 71T T8E BELL$VOE PI~RE 8IT8. BT. MARYB RIVER
Unlike Hamilton Harbour, metals (Figure 5) and the 16
priority pollutant polynuclear aromatic hydrocarbons (PAHs, Table
1) are comparatively dilute at the Bellevue Park test site in the
St. Marys River (Figure 6). However, sediment samples from near
Bellevue Park have high concentrations of oil and grease (1.4%,
1.6% and 2.4%) and wood fibres.
A high concentration of a complex PAH (retene,~ 2 ~tg/g) was
found in the St. Marys River sediments. Retene can occur
naturally from degradation of conifers but it can also be
associated with pulp and paper manufacturing. The concentrations
1~ _~~~~4~
of many of the priority pollutant chlorinated organic compounds
is near or at background levels (Tables 2 and 3).
Decay of the wood fibre and other wastes results in a
reducing environment as indicated by the black colour, high
ammonia (1.5-2.3 mg/L), and low redox (Figure 7). Note that the
deeper sediments are more oxic. This observation reflects the
relatively recent discharge of labile organic wastes over older
more oxic sediments. Also note that the redox of the sediments
l0 changes seasonally. By November the surface sediments have
become oxygenated (Figure 7). It is a fortunate situation in
that oxidation treatment of the surface sediments could not be
compromised by diffusion of reduced materials such as hydrogen
sulphide from deeper sediments. Also the required depth of
treatment is only 15 cm.
18
Table 1
PAH Concentrations in Surficial Sediment
of Bellevue Marine Park Area
Sample Range
(ng/g)
Naphthalene-' 3137 6878
-
Acenaphthylene 152 - 318
Acenapththalene 169 - 360
Fluorene 356 - 540
Anthracene 1913 3425
-
Phenanthrene 478 - 1227
Fluoranthene 2599 6831
-
Pyrene 2021 5485
-
Chrysene 1068 3269
-
Benzo(a)anthracene 1353 3680
-
Benzo(b)fluoranthene 2004 2223
-
Benzo(k)fluoranthene 1512 2202
-
Benzo(a)pyrene 964 - 3114
Dibenzo(a,h)anthracene 275 - 1040
Inden~ (1 , 2, 3-cd) i3u - 41i
rl~re.~.~
Benzo(g,h,i)perylene 370 - 1214
Total PAHs 16989 42019
-
19
Table 2
PCB Concentrations in Surficial Sediment
of Bellevue Marine Park Area
Sample Range
PCB 18 ND 0.66
PC-8-52 ND - S . 13
PCB 49 ND - 5.43
PCB 44 ND - 3.69
PCB 101 2.48 - 5.77
PCB 151 ND - 9.23
PCB 118+149 2.29 - 15.57
PCB 105 ND - 1.92
PCB 138 2.91 - 6.36
PCB 183 ND - 2.94
PCB 194 ND - 0.85
Total 80.43 - 299.28
PCB
20
Table 3
Organic Contaminant Concentrations in Surficial Sediment
Of Bellevue Marine Park Area
Sample Range Sample Range
(ng/g) (ng/g)
1,3 DCB - -- -ND Aldrin ND
1,4 DCB ND OCS ND
1,2 DCB ND g Chlordane ND
HCE ND o,p DDE ND
1,3,5 TCB ND a Endosulfan ND
1,2,9 TCB ND a Chlordane ND
1,2,3 TCB ND t Nonachlor ND
1,2,3,5 TECB ND Dieldrin ND
1,2,4,5 TECB ND p,p' DDE 1.97-3.17
1,2,3,4 TECB ND ~ o,p' DDD ND
PECB ND Endrin ND
2,3,4,6 TECB ND B Endosulfan ND
A BHC 1.66-4.14 p,p' DDD 1.22-3.32
HCB ND o,p' DDD ND
PECA :~;D liCl.llC!?CYV.i:iv
Lindane ND Mirex ND
Heptachlor ND
21
The highest observed concentration of hydrogen sulphide in
the sediment was in June (Figure 8j. By the end of August most
of the hydrogen sulphide had been oxidized. The sediment
treatment with ferric chloride greatly reduced the concentration
of hydrogen sulphide.
The ATP-TOX bioassay indicated a seasonal change in toxicity
(Figure 9) that closely matched that of the hydrogen sulphide
concentration. Also, the ferric chloride treatment reduced the
toxicity of the ATP-TOX bioassays in tandem with the hydrogen
sulphide complexation. Photobacteriuai~hosphoreum bioassays also
indicated a seasonal change in toxicity but they were not done
as intensively as the ATP-TOX bioassays (Figure 10). Daphnia
magna bioassays with aqueous extracts indicated no toxicity
(Figure 11 but some DMSO extracts with Daphnia could measure
toxicity (Figure 12).
Two other bioassays indicated little or no toxicity.
Dilution bioassays with Hexaaenia (mayfly nymphs) from four field
trips in 1991 (February, June, July, and August) indicated no
toxicity. Bioassays with Lactuca 'va (lettuce) detected
little toxicity. These latter bioassays are not sensitive to
hydrogen sulphide toxicity.
The bioassay results were not as clear as visual inspection
of the site was. Very few benthic invertebrates were seen: the
sediments were virtually sterile to organisms. The main acute
toxin is hydrogen sulphide and some laboratory bioassays cannot
detect this toxin.
Three sets of ~aphn~ toxicity dilution experiments in
Hamilton Harbour also observed a seasonal change in sediment
toxicity (Figures 13, 14 and 15). Some variation exists and the
trends are more obvious by looking at the average toxicity
(Figure 16). In winter, these sediments have little toxicity.
However, if these winter samples are purged with nitrogen, then
sealed for a month to go anoxic, then bubbled with air for 2-3
22
h to oxygenate them, they are highly toxic. This length of
oxygenation provides oxygen saturation but it is less than the
half life of hydrogen sulphide oxidation (19 h).
By late fall, the sediment samples from the Stelco Hotspot
were still highly toxic. These observations differ from the St.
Marys River sediments where hydrogen sulphide toxicity was almost
gone by late August. The differences in extremes of hydrogen
sulphide concentrations support the hypothesis that the Stelco
Hotspot with 100 mg/L of hydrogen sulphide will stay toxic for
much longer than the St. Marys River sediments with 4 mg/L
hydrogen sulphide.
New toxicity maps with Daphnia na (Figure 17) and
Photobacterium phosphoreum (Figure 18) indicate much less
toxicity than earlier maps (Brouwer et al. 1990). The new maps
are done from analyses of surface sediment (0-1 cm), whereas the
old maps were done from analyses of Ekman dredge samples (0-15
cm). In part, the surface sediments have less contaminants, but
the deeper sediments have less access to oxygen and anoxic decay
produces hydrogen sulphide.
If the acute toxicity is controlled by hydrogen sulphide,
then no biodegradation occurs in the deeper sediments and some
occurs at a suppressed rate in the most recent sediments. The
proof of this last hypothesis is found in the PAH data. The
surface sediments have much less naphthalene than the deeper
sediments (Tables 4 and 5). Naphthalene is biodegradable (Murphy
et al. 1992). Some recovery from source control is occurring,
but it is slow. The rate of recovery is uncertain in the deep
basin. The deeper sediments of the Stelco Hotspot, i.e., >2 cm
deep are not recovering and will likely remain uninhabitable to
benthos for decades.
Two sediment injection trials were conducted in the St.
Marys River near Bellevue Park. The first trial was relatively
successful but modifications were made to improve efficiency
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Table 6
Summary of Sediment Trap Data - St. Marys River Chemical Treatments
Sediment Traps (g) Sediment Traps (g)
July 8, 1991 Oct 10, 1991
Upstream Downstream Upstream Downstream
0.122 0.182 0.066 0.049
0.140 ' O.r69 0.064 0.029
0.154 0.144 0.097 0.074
0.081 0.158 0.048 0.081
0.497 0.653 Total 0.225 0.233
0.124 0.163 Average 0.056 0.058
+31.4 ~ Dif +3.6
Surface Water (g/L) Surface Water (g/L)
July 8, 1991 Oct 10, 1991
Upstream Downstream Upstream Downstream
0.003 0.009 0.0017 0.0015
28
before the second injection trial. The system, described herein
previously, for injecting iron into sediments was built and
tested for the first time in the St. Marys River on July 10,
1991. The system had the capacity to treat a large area: three
45-gallon barrels of ferric chloride were injected in less than
an hour over an area 90 m by 12 m. The equipment was tough in
that it survived bouncing over logs and other debris. In spite
of minor engineering problems, the equipment worked well. The
colour of the iron in the sediment cores indicated that the top
9 cm of sediments were treated. Refinements in the pumping
equipment were needed before larger areas could be treated. The
chemical pump required an air compressor and only large ones were
available from rental agencies. A compact Honda air compressor
was bought to provide more free deck space. The pore size of the
nozzles was large: to pressurize the manifold to achieve equal
flow through all nozzles, the flow rate of ferric chloride was
higher than planned.
Divers recorded with cameras that the sediments were not
resuspended into the water column. A pressure wave proceeding
the injection bar raised the sediments about 20 cm, but they fell
back with minimal resuspension to the upper waters. Sediment
trap analysis confirms that no sediment moved to the water
surface (Table 6). A large amount of gas reached the surface and
small patches of oily film formed. Some macrophytes (primarily
E1 ea canadensis) were broken by the injection bar, but most
remained intact. The sediments need to be studied in more detail
but the ATP-TOX bioassay indicated a reduction in acute toxicity
after the ferric chloride treatment (Figure 9).
On October 6, a second ferric chloride injection was done
north of the first site to an area 200 m by 36 m. Smaller
nozzles (0.031 inch diameter orifice) were used to maintain a
high back pressure in the injection manifold and a constant flow
through all the nozzles. The skids on the injection bar were not
used so that the injection bar could penetrate deeper into the
sediments. Visual observations indicated that the surface 15 cm
,~...r.. 2099423
29
of sediments were treated. Observations that with the boom
configuration, the boat could not move faster than 0.5 m/s or the
injection manifold would rise above the sediments.
$Z~MPLg Z - CAhCIO~I ~TITRATB INJECTION
Calcium nitrate is about 100,000 times more water soluble
than oxygen. The following reaction is mediated by bacteria:
5CH0z0 + 4N03- - 2N2 + 4HC03- + C02 + 3H20.
Surface sediments (1-50 cm) from the St. Marys River were
collected near Bellevue Park as illustrated in Figure 6. Surface
sediments (0-15 cm) were collected from the deep basin of
Hamilton Harbour as illustrated in Figure 19. Sediments were
collected with either a Shipek dredge, or a Tech Ops corer. In
Hamilton, samples were returned to NWRI and placed in either a
fridge or 12°C incubator. In Sault Ste. Marie, samples were
quickly placed in a cooler and stored in a fridge in the Great
Lakes Forestry Institute. Samples were always placed in coolers
and shipped quickly. Sediment cores were subdivided within 24
h. Sediment samples were placed in clean pails with lids and
enough sediment was added to exclude air. Sample processing for
bioassays included homogenization with larger mixer and
subsequent handling in a glove box in a fumehood. Sampling of
reactors was done in a glovebox after purging it with nitrogen.
Various bottles were tried. The first trials used 300 mL
BOD bottles to incubate and measure microbial utilization of
nitrate. Each sample was unique in that after opening the top,
the sample was not reincubated. The microbial metabolism
incubations were run using 300 mL BOD bottles with and without
100 mg N/L of calcium nitrate. The short-term experiments were
successful, but the production of gas ruined longer incubations
by popping the lids. Both 155 mL and 250 mL septum fitted
bottles were used in subsequent trials for incubations with 500
mg/L N-N03. Biodegradation experiments were run in 155 mL glass
bottles with serum caps and a 20 mL nitrogen headspace, with and
X099423
without 500 mg/L N calcium nitrate. The nitrogen headspace was
sampled with gas tight syringes after relatively short-term
incubations (2-6 weeks).
5 For all incubations, the sediments from the St. Marys River
were mixed with deoxygenated water from St. Marys River and the
Hamilton Harbour sediments were mixed with dechlorinated
deoxygenated Burlington City water to form a 50% slurry. All the
above sediments were shaken continuously on an end-over-end
10 shaker. In one trial to measure the production of ammonia, 2 L
jars were used for incubations with 500 mg/L N-N03; these
sediments were shaken once a day (except some weekends).
The sediment slurry was centrifuged and the supernatant was
15 filtered and processed using an ion chromatograph to determine
nitrate and sulphate concentrations. The pH of samples was
measured with a pH meter. Ammonia was analyzed by colorimetric
analysis (Solorzano 1969).
20 Volatile organic compounds in the headspace were measured
by GC/MS in the Waste Water Treatment Centre (WTC) laboratory
(Brian MacGillivray). Each assay was processed with five
replicate bottle incubations and the headspace subsamples were
combined. For one experiment two sets of five replicates were
25 processed to determine the analytical error; it was
insignificant. Sediment samples for hydrogen sulphide analyses
were frozen and delivered to Guelph Chemical Laboratories. These
samples were purged with helium without any pH treatment; the
hydrogen sulphide was trapped in a cold trap and injected for
30 analysis into a GC/MS.
"Oil and grease and total petroleum hydrocarbons" was
measured with a derivative of the Environment Canada (1979)
protocol. "Total petroleum hydrocarbon" was measured by a gas
chromatography method. Sodium sulphate was used to dry the
samples. Dichloromethane was used to extract the samples with
209923
31
cycles/h in a Soxhlet extractor for 8 h with a water bath at
25°C.
~hotobacterium bioassays were run on whole sediments
5 (Brouwer et al. 1990). Dilutions for LCSO analysis were done
with clean sediments from Long Point, Lake Erie.
Results
Enhancement of Microbial Metabolism
10 St. Marys River Sediments.
The initial laboratory reactor experiments have been
successful in stimulating microbial metabolism in St. Marys River
sediments with calcium nitrate. The microbial denitrification
of nitrate was coupled to the rapid production of sulphate, this
is illustrated in Figures 20 and 21. The production of sulphate
reflects the microbial oxidation of organic sulphur, hydrogen
sulphide, and perhaps elemental sulphur. The 100 mg/L N-N03 dose
was completely denitrified (Fig. 20). The next experiment with
500 mg/L N-NO3 resulted in incomplete denitrification of the
added nitrate (Fig. 21) . This sample was collected in late fall;
the longer lag phase in the second experiment probably indicates
that the microbes were inactive and needed more time to produce
enzymes.
After two weeks of incubation, sediments treated with 500
mg/L N-N03 were given to Dr. Reynoldson (NWRI) for bioassays.
They were toxic to Hexagenia . Based upon results from the second
experiment (Fig. 22), these sediments had a high concentration
of nitrate (>300 mg/L N-N03). Either the nitrate caused osmotic
shock or the intense production of nitrogen gas disrupted their
intestines. This experiment needs to be redone with less calcium
nitrate. There would be no similar negative response in situ
because no benthos would be living in sediments requiring
treatment. However, this experiment does illustrate the need to
balance the treatment dose to the biological oxygen demand of the
sediments.
2099423
32
In another trial using larger containers (2 L) without
continuous shaking, the utilization of nitrate was slightly
slower than in Figs. 21 and 22). Since these incubations used
50% slurries, these incubations indicate that the optimal in situ
does is about 350 mg/L N-N03. Very little ammonia was produced
during these incubations (Figs. 23-24). Other short-term trials
indicated that phosphorus was not limiting microbial
denitrification, that addition of iron did not suppress
denitrification, and that the pH did not decrease significantly.
For short-term incubations, nutrients did not appear to limit
microbial metabolism and pH buffering was not required.
In one year long incubations with sediments from the St.
Marys River, nitrate treatment resulted in biodegradation of
about 60% of the polynuclear aromatic hydrocarbons (PAHs, Figure
25). The numbers in Figure 25 refer to the molecular weight of
the 16 priority pollutant polynuclear aromatic hydrocarbons.
Note that the larger molecular weight compounds that can induce
tumours are biodegraded as well as the smaller compounds. This
observation is inconsistent with radioisotope studies done in our
laboratory and in other laboratories. Smaller radiolabelled PAHs
can biodegrade very quickly (within weeks). As observed in the
Hamilton harbour studies, the conclusion is that very large PANS
that are too large to measure with available technology are
biodegrading to produce measurable PAHs which continue to
biodegrade. The rate limiting step on the treatment is the
biodegradation of the large PAHs and the treated sites will
require about two years for effective PAH treatment.
Hamilton Harbour
The rate of denitrification in a sediment sample from the
deep basin of Hamilton Harbour was slightly slower than in a
sample from the St. Marys River (Bellevue site), and much slower
than in a sample from the Stelco Hotspot (Fig. 26). In the
sample from the deep basin of Hamilton Harbour, the
denitrification resulted in the complete elimination of toxicity
to Photobacterium (Fig. 27). The incubated sample from this
33
Stelco Hotspot had residual hydrogen sulphide and toxicity
persisted. More nitrate has been added to the Stelco Hotspot
incubation and analysis will be replaced after a further six week
incubation.
The oxidation of the deep basin sediments is more obvious
than other sediments. These sediments were black at the start
of the incubations, the control samples remained black, but the
calcium nitrate treated samples turned brown. The sediments from
the St. Marys River were not as black, but in long incubations
(three months), the control samples turned black and the calcium
nitrate samples stayed brown. The colour change reflects the
change in the oxidation state of iron.
Another simple physical change also occurs during
treatments. The control samples are very flocculant and these
sediments stay in suspension for days. The treated sediments are
not flocculant; these sediments precipitate within three hours
after shaking. The treatments must polymerize negatively charged
organic colloids. This flocculation could be very useful.
Colloids can contain high concentrations of contaminants and
their resuspension in dredging can create problems.
In general, the NWRI studies are consistent with published
demonstrations of biodegradation but the processes are complex
and this study is not complete. Simple analyses of "oil and
grease" indicate about 50% biodegradation of the organic
contamination. However, "total petroleum hydrocarbon" analysis
indicate about 90% biodegradation of organic contamination. Some
of the discrepancy is likely caused by the microbial conversion
of organic contaminants into organic compounds in living cells.
The PAH data from headspace analysis is complex but highly
encouraging. Headspace analysis of some samples indicates
biodegradation of butenes, chlorobenzenes, toluene, benzene, and
naphthalene at rates consistent with other studies. Other
analyses indicate production of several compounds, indicating
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36
cleavage of smaller molecular weight compounds from larger
compounds. There are about 30 analyses as complex as Figure 28.
Further synthesis is required to resolve the complexities. Note
that the analyses were replicated and the analytical error was
insignificant.
The PAH analysis of the solids remaining after six weeks
incubation with or without calcium nitrate indicated no
significant biodegradation of naphthalene or other PAHs (Table
7): Analytically, the discrepancy with headspace analysis is
possible in that the headspace represents only a small fraction
of the total PAHs. The headspace is in equilibrium with free,
unbound, bioavailable compounds, but the particulate PAH analysis
is done on samples extracted vigorously with dichloromethane in
a Soxhlet apparatus. There are two hypotheses that can resolve
the PAH data sets. 1) Some small compounds were being cleaved
from larger compounds at similar rates that microbes were
biodegrading these smaller organic compounds. 2) Only a small
fraction of the PAHs are biodegradable.
Both hypotheses could be valid. Only a fraction of the
total PAHs are analyzed in the routine 16 priority PAHs, no
analytical techniques exist for very large molecular weight PAHs.
Also for both sites there must be some PAHs locked in coal dust
or other biologically inactive matrices. The interpretation
would also vary between sites. For example, the proportion of
PAHs in the aqueous phase in the St. Marys River incubations
indicates enhanced production of naphthalene from larger
compounds after the treatments (Table 8). In the aqueous phase
of Hamilton Harbour incubations, the treatments appear to enhance
biodegradation of naphthalene (Table 8). The uncertainties of
PAH biodegradation have been resolved with longer incubations,
and incubations with '4C-radioactively labelled naphthalene.
Pilot-scale treatments also support the laboratory
incubations. The addition of calcium nitrate to the sediments
of the Dofasco Boatslip in 1992 resulted in the biodegradation
2~9g4~3
37
of several organic compounds (mean of three samples, reductions
as follows; toluene 80%, ethylbenzene 86%, m/p-xylene 76%, 3/4-
ethyltoluene 89%, and dichloromethane 65%) (Figure 29). These
relatively rapid biodegradation rates are similar to those
reported in laboratory studies where nitrate was added to enhance
biodegradation (Hutchins 1991).
Analysis of three samples indicates that 25% of the
petroleum hydrocarbons were biodegraded in the Dofasco boatslip
l0 treatment.
The biodegradation of the PAHs (polynuclear aromatic
hydrocarbons), in the Dofasco boatslip was more complex. About
15% (450 ~Cg/g, to 383 ~g/g mean of 3 samples) of 15 PAHs were
biodegraded and in the process the naphthalene content increased
196% (280 ~g/g to 549 ~,g/g, mean of 3 samples). The imbalance
in the concentration of naphthalene suggests that other higher
molecular weight compounds not measured in the standard priority
pollutant PAH analysis are decomposing to produce naphthalene.
Approximately 50% of the PAHs in coal tar pitch contain more than
seven rings (Enzminger and Ahlert 1987); we are capable of
measuring less than 50% of the PAHs.
At first the ability of microbes to biodegrade organic
wastes seemed less probable in Hamilton Harbour sediments than
in sediments from the St. Marys River. Hamilton Harbour
sediments have 10-100 times the concentration of metals.
However, the rate of headspace naphthalene biodegradation is
similar in sediments from Hamilton Harbour, St. Marys River, and
samples from other sites (Heitkamp and Cerniglia 1987).
Furthermore, the rates of denitrification in the St. Marys River
sediments, Hamilton Harbour, and other sites in Germany (Ripl
1986) are similar. At these sites, metals do not appear to
suppress microbial biodegradation. This is important in that
35, many of the volatile organic compounds that were detected in the
Hamilton Harbour Hotspot (Table 9) are biodegradable.
38 2099423
I?I8CU88ION
The oxidation of toxic hydrogen sulphide eliminates most of
the acute toxicity from Hamilton Harbour and St. Marys River
sediments (Murphy et al. 1992). The extreme anoxia reflected by
high concentrations of hydrogen sulphide inhibits microbial
biodegradation. In headspace analysis, some simple compounds
like butene, naphthalene, and toluene appear to be biodegraded
within weeks of nitrate addition. The biodegradation of larger
non-volatile organic contaminants such as benzo(a)pyrene will be
slower, albeit the published rates vary greatly. Heitkamp and
Cerniglia (1987) found that naphthalene, pyrene and
benzo(a)pyrene would degrade with half-lives 1.4-4.4 weeks, 38-90
weeks and 200-300 weeks, respectively. Shiaris (1989) found
biodegradation turnover times of 13.2-20.1 days, 7.9-19.8 days,
and 53.7-82.3 days for naphthalene, phenanthrene, and
benzo(a)pyrene, respectively. Ongoing long-term bioassays in
NWRI will help resolve the biodegradation of larger compounds
like benzo(a)pyrene. The optimal study, however, would be the
monitoring of pilot-scale applications of calcium nitrate to the
sediments of as many sites as possible. Each site will be
slightly different and new insights will develop from each
treatment.
COMP~IRIBa~1 pITB FBRRIC CBI~O~RIDB TREATMBNT
Because of the engineering success of the related
experiments done by NWRI (Murphy et al. 1992), the calcium
nitrate treatments have quickly become pilot-scale treatments.
Both ferric chloride and calcium nitrate are oxidants. Ferric
chloride is a weaker oxidant, albeit is reactions with metals and
hydrogen sulphide are more direct and potentially useful. To
achieve the equivalent oxidation potential of a 0.5% solution of
calcium nitrate would require that the sediments become a 10%
ferric chloride solution. This latter scenario would produce a
toxic low pH that would require extensive buffering with lime.
Moreover, calcium nitrate is less corrosive to equipment than
ferric chloride. The chemical cost of treating the surface 15
cm of sediments with 500 mg/L N-N03 would cost $2,000 to $10,000
a hectare. The range of costs reflects the chemical oxygen. The
39
''~ zfl99423
Table 9
STELCO HOTSPOT SEDIMENT
ANALYSIS BY PURGE AND TRAP GC/MS
PARAMETER ng/ml
1,1-dichloroethylene 659.2
dichloromethane 14.3
trans-1,2-dichloroethylene11.3
1,1-dichloroethane 97.2
cis-1,2-dichloroethylene0.0
chloroform 13.0
1,1,1-trichloroethane 0.0
tetrachloromethane 0.0
1,2-dichloroethane 18.1
benzene 831.2
trichloroethylene 0.0
1,2-dichloropropane 0.0
dibromomethane 0.0
bromodichloromethane 0.0
toluene 596.8
1,1,2-trichloroethane 0.0
tetrachloroethylene 21.6
chlorodibromomethane 0.0
1,2-dibromoethane 0.0
chlorobenzene 0.0
~
ethylbenzene 1348.8
m/p-xylene 3002.0
o-xylene 1225:2
styrene 279.3
cumene (isopropylbenzene)119.3
bromoform 0.0
1,1,2,2-tetrachloroethane5.9
propylbenzene 112.9
1,3,5-trimethylbenzene150.8
1,2,4-trimethylbenzene14.1
3-ethyltoluene 1050.0
4-ethyltoluene 1247.2
2-ethyltoluene 1598.0
1,3-dichlorobenzene 10.3
1,4-dichlorobenzene 0.0
1,2-dichlorobenzene 0.0
1,9-diethylbenzene 191.7
1,2-diethylbenzene 9.3
1,3-diethylbenzene 191.7
naphthalene 35920.0
hydrogen sulphide 100000.0
4~ ~099~~3
Table 10
CHLORINATED PHENOLS IN ST. MARYS RIVER
SEDIMENT CORE - JULY 8, 1991
SAMPLE ID (CORE DEPTH - cm) 0-1 6-7 12-14 19-16 16-18 22-24
Chlorinated Phenols
(ug/kg) Dry Weight
ortho-Chloro-pheno l
meta-Chloro-phenol
para-Chloro-phenol
2,6-Chloro-phenol ND ND ND ND ND ND
2,9-Chloro-phenol ND ND ND ND ND ND
3,5-Chloro-phenol ND ND ND ND ND ND
2,3-Chloro-phenol ND ND ND ND ND ND
3,4-Chloro-phenol ND ND ND ND ND ND
2,4,6-Chloro-phenol ND ND ND ND ND ND
2,3,6-Chloro-phenol ND ND ND ND ND ND
2,3,5-Chloro-phenol ND ND ND ND ND ND
2,4,5-Chloro-phenol ND ND ND ND ND ND
3,4,5-Ghloro-phenol ND ND ND ND ND ND
2,3,5,6-Chloro-phenol ND ND ND ND ND ND
2,3,4,5-Chloro-phenol ND ND ND ND ND ND
Penta-Chloro-phenol ND ND ND ND ND ND
4-Chloro-guaiacol
4,6-Chloro-guaiacol ND ND ND ND ND ND
4,5-Chloro-guaiacol ND ND ND ND ND ND
3,4,5-Chloro-guaiacol ND ND ND ND ND ND
4,5,6-Chloro-guaiacol ND ND ND ND ND ND
3,4,5,6-Chloro-guaiacol ND. ND ND ND ND ND
4-Chloro-catechol
3,5-Chloro-catechol** ND ND ND ND ND ND
2,3,4,6-Chlorophenol**
4,5-Chloro-catechol ND ND ND ND ND ND
3,4,5-Chloro-catechol ND ND ND ND ND ND
3,4,5,6-Chloro-catechol ND ND ND ND ND ND
6-Chloro-vanillin NA NA NA NA NA NA
5,6-Chloro-vanillin ND ND ND ND ND ND
Tri-chloro-syringol ND ND ND ND ND ND
4,5 Di-chloro-veratrole ND ND ND ND ND ND
3,9,5 Tri-chloro-veratrole ND ND ND ND ND ND
Tetra-chloro-veratrole ND ND ND ND ND ND
GUGICOL
CATECOL *
* analysis for these compounds to follow
** these compounds coelute
ND not detected
NA not applicable
41 zo~o423
sediments of the St. Marys River and deep basin of Hamilton
Harbour require less than a third of the dose required for the
Stelco Hotspot. With the 8 m injection boom and ideal
conditions, about four hectares a day could be treated.
The sediments of the St. Marys River appear to be easy to
treat. The acute toxicity is caused by hydrogen sulphide and it
is readily oxidized by denitrification of added calcium nitrate
(Murphy et al . 1992 ) . The oxidized sediments produce no toxicity
l0 to Daphnia magna, Hexaqenia limbata, Bscherichia coli, or Lactuca
sativa. Many chlorinated compounds often associated with
pulpmill wastes were not detected (Table 10). Benzenes are
detectable in these sediments and although they are carcinogens,
they are biodegradable.
The concentration of PAHs is relatively low (Table 3: Murphy
et al. 192, Table 1). The concentration of PAHs is approximately
at the apparent effects threshold - the concentration where you
begin to see toxic effects on the ecosystem (Long and Morgan
1990). There is no evidence that the PAHs in the sediments at
the Bellevue site are a problem. Chemical data, particularly
threshold concentrations, must be used cautiously. The high
concentration of organic matter found in the sediments of the St.
Marys River could reduce the bioavailability of PAHs as has been
found at other sites (Landrum et al. 1987). As well as
conducting additional chemical analyses, the endpoint of the
biodegradation should be determined with bioassays. The best
bioassay would be the response of the benthos in treated
sediments.
