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PRIMARY TREATMENT
INTRODUCTION
Primary treatment is intended to physically remove settle-
able solids and most of the discrete suspended and floating
solids from the municipal wastewater stream preparatory to
secondary treatment. In addition, primary treatment removes a
limited portion of the soluble constituents.
In primary treatment, the wastewater influent is divided
into three output pathways: primary effluent, primary sludge
(including grit, screenings, and precipitated matter), and
aerosols. Effluent from primary treatment can be directly
discharged (until 1977), discharged after disinfection (again
until 1977), or treated by a secondary process. Primary sludge
is normally subject to additional processing. At some ocean
coastal sites, however, the sludge is discharged without further
treatment. Aerosols are rarely a problem, due to the absence
of excessive turbulence.
The information reviewed during this study is tabulated in
Table 12. Most research conducted to date concerns the removal
by primary treatment of water quality parameter constituents.
It is only in recent years that researchers have examined the
effect of primary treatment on various public health impairing
contami nants.
WATER QUALITY PARAMETERS
For several decades, extensive work has been reported con-
cerning general contaminant removal efficiencies for primary
treatment. Removal varies widely, depending upon the physical
and chemical characteristics of the wastewater, the proportion
of settleable solids, concentrations of the solids, and deten-
tion time. Mi tchel 1 (922) reported the general performance effi-
ciencies that could be expected for typical primary treatment
(Table 13). Removals achieved during a three-year study of
an operational primary treatment system are shown in Table 13.
78
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TABLE 12. LITERATURE REVIEWED PERTAINING
TO PRIMARY WASTEWATER TREATMENT
Contaminant Reference Number
Water Quality Parameters
Ammonia 922
BOD 393, 622, 745, 922, 1404, 1435, 1547
COD 483, 622, 896, 922, 1435
Cyanides 922
Oil and grease 483, 810, 922
Phosphates 896
Suspended solids 393, 483, 622, 745, 896, 922, 1435, 1547
Total organic 483
carbon
Other (general) 531, 622
Elemental Contaminants
Arsenic 922
Cadmium 228, 443, 922, 1063
Chromium 228, 896, 922
Copper 228, 443, 896, 922, 1063
Iron 228, 896
Lead 228, 922, 1063
Manganese 228
Mercury 228, 922
Nickel 228, 922, 1063
Selenium 1114
Zinc 228, 896, 922, 1063
79
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TABLE 12 (continued)
Contaminant Reference Number
Biological Contaminants
Bacteria 79, 161, 1256
Coliforms 161, 483, 568, 717, 896, 1256, 1282
Coxsackie virus 161
(A & B)
Escherichia coli 568, 896
Streptococci 568
Hepatitis virus 1261
Mycobacterium 161, 428
Parasitic worms 161, 420, 428
Pol io vi rus 97, 161
Protozoa 161, 428
Salmonella 161, 420, 428, 568, 1261
Shigella 161, 1261
Vibrio cholerae 161
Virus 79, 92, 95, 96, 97, 100, 101, 161, 382,
428, 468, 1009, 1256
Other (general) 161, 1261
80
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TABLE 13. PRIMARY SETTLING TANK PERFORMANCE (922)
Parameter
COD
BOD5
Suspended Solids
Oil and Grease
Ammonia Nitrogen
Primary
Influent
(mg/£)
539
269
279
72
34
Primary
Effluent
(mg/£)
315
165
103
28
20
Percent
Removal
42
39
64
61
41
When used as a coagulant during primary treatment, lime is
most effective in reducing certain water quality parameter con-
stituents. When lime addition of 350 mg/£ was followed by air
flotation, Mennell et al . (896) reported the following removals:
turbidity, 98.5 percent; suspended solids, 95 percent; COD, 60
percent; and total phosphorus, 99 percent. Total nitrogen
removal varied between 10 and 20 percent. Lower lime dosages
provided proportionally lower removal percentages. The practice
of dosing primary clarifiers with chemical coagulants will
probably increase, as municipalities attempt to cost effectively
meet federal and state water quality standards.
ELEMENTAL CONTAMINANTS
Recent interest in elemental contaminants, particularly
trace metals, has prompted investigation into the partitioning
of these contaminants in the wastewater treatment stream.
Primary treatment receives elemental contaminants in a variety
of forms, e.g., soluble, insoluble, and complexed. The con-
centrations of each form vary intrinsically as a complex
function of such factors as influent metal concentration, pH,
and ligand concentration.
Mitchell (922) reported removal efficiencies for primary
settling over a three-year period at the Hyperion treatment
plant, Los Angeles, California,as shown in Table 14.
TABLE 14. PRIMARY TREATMENT REMOVAL OF METAL ELEMENTS (922)
Copper
Zinc
Nickel
Lead
Arsenic
Cadmium
Chromi urn
Primary
Influent(mg/£)
0.39
0.66
0.30
Q.03
0.015
0.01
0.55
Primary
Effluent(mg/£)
0.25
0.42
0.24
0.07
0.017
0.02
0.37
Percent
Removal
36
36
20
32
81
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No explanation was offered to account for the anomaly of
increased lead, arsenic, and cadmium concentrations.
Chen and Lockwood (228), also working with the Hyperion treatment
plant, discussed the partitioning of trace metals with particu-
lates as a function of particle size. It was reported that in
primary effluent, more than 70 percent of the total trace metals
content is associated with particulates as opposed to the solu-
ble ionic form. However, only 10 to 20 percent of the Ni, Pb,
or Mn was associated with the particulate fraction. Similar
concentrations of metals were found on the larger (44 ym) and
the smaller (0.2 ym) particles. However, particles as large as
44 ym will settle about 200 times as rapidly as 3-ym particles.
More efficient removals of elemental contaminants could be
expected by the use of either increased wastewater detention
time, or chemical precipitants (e.g., lime) for particle coagu-
lation. Trace metal removals reported (896) at a lime dose of
388 mg/£ approached 100 percent for chromium and copper, 97
percent for iron, and 94 percent for zinc. Molybdenum was not
effectively removed by this treatment.
BIOLOGICAL CONTAMINANTS
Primary sedimentation usually removes less than 50 percent
of coliform and pathogenic bacteria from sewage and is rela-
tively ineffective in removing viruses and protozoa. Literature
concerning the removal of water-borne pathogens by primary
treatment processes reports a varying degree of efficiency,
depending in part on the type of pathogen studied. Table 15
(161, 428) highlights the results of Bryan's investigation of
pathogen survival during primary treatment.
In their literature review, Foster and Englebrecht (428)
reported isolation of salmonella from six of seven different
primary effluent samples. The raw sludge also contained members
of this genus, with 19 of 20 samples tested as positive for
salmonella organisms. Tubercle bacilli were reduced about 50
percent in the wastewater stream during sedimentation. It was
concluded that bacterial pathogens are ineffectively removed
from wastewater by primary settling; furthermore, the process
produced a sludge that, without further treatment, constitutes
a health hazard.
Amoebic cysts and parasitic worm ova are also ineffectively
removed by primary treatment, due to their low specific densities
and resulting buoyancy. Ascaris ova are an exception: Foster
and Engelbrecht (428) reported 100 percent settlement of these
ova into the sludge within 15 min.
Viruses, because of their size (,02 to .3 ym), are rarely
removed by sedimentation, except for those that are associated
with wastewater solids. The nature of the surface chemistry of
82
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TABLE 15. SURVIVAL OF PATHOGENS
DURING PRIMARY TREATMENT (161, 428)
Pathogen
Salmonella typhi
Salmonella spp.
Streptococcus faecal is
Mycobacterium
Enteroviruses
Polio viruses
Coxsackie viruses
Endamoeba histolytica
Ascaris ova
Trichuris trichiura
Taenia saginata
% Removal
>50
0-15
<50
48-57
no reduction
no reduction
<50
0-50
0-100
>50
0-50
83
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viral units suggests that adsorption depends strongly on the
chemical environment, would be expected to vary according to
changes in input water chemistry, and would possibly be affected
by chemical additions.
The removal of viruses by primary settling has been both
researched and reviewed extensively by Berg (92, 95, 97, 100).
Berg described a project (95) in which only one to two-thirds
of the viral particles had settled out in one day, although 75
percent of the suspended solids had settled. In this review,
Berg discussed several additional studies on viral removal by
primary treatment; all failed to mention detention time, and
more importantly, the levels of virus in the incoming sewage
were not related to those in the effluent.
Although primary settling alone will not effectively reduce
the pathogen content of wastewater, dosing primary settling tanks
with chemical coagulants does show some promise in this regard.
Chemical precipitation, when used during primary treatment, is
capable of removing as much as 99.99 percent of the virus sus-
pended in water, effected through the formation of a coagulant-
cation-virus complex. Elevated pH levels attained during lime
treatment also result in substantial reductions in viral numbers
(468). Lime coagulation during primary treatment brings remark-
able reductions of coliform density as well. A 99.9 percent
coliform reduction was measured at a lime dose of 450 mg/£ (896).
84
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SECONDARY TREATMENT: ACTIVATED SLUDGE
INTRODUCTION
The activated sludge process involves the growth of
microorganisms in a reactor. This effects partial biological
degradation of organic compounds in wastewater to simpler
organic compounds, carbon dioxide, water, microorganisms, and
energy (393). The basic process requires two equipment compo-
nents: aeration tanks and clarifiers. Active biological sludge
is separated from the effluent in a clarifier and recycled to an
aeration tank.
Activated sludge, the most popular wastewater secondary
treatment process, has been extensively studied, as indicated
by Table 16 . Most research has focused on water quality
parameters such as BOD, COD, and suspended solids. The data
are usually presented as percent removal, with removal effi-
ciency determined by difference in influent and effluent con-
centrations. Removal of a specific contaminant can be
accomplished by separation into the sludge or by degradation
through biological activity. Aerosol generation from the
aeration tank is also a possible contaminant pathway.
In view of possible health effects, the difference between
separation and degradation can be significant. If the treatment
process merely partitions a particular contaminant into the
sludge or air, it remains available for migration back to man.
In contrast, biological degradation can terminate the contami-
nant pathway or transform the potentially harmful substance into
a nontoxic form. The separation and degradation components of
the removal process are often not distinguished in the activated
sludge literature.
WATER QUALITY PARAMETERS
Past research on activated sludge processes has concen-
trated on water quality parameters, with primary emphasis on
BOD, COD, and suspended solids. Because of the tremendous
volume of literature associated with BOD and suspended solids as
indicators of activity, and the absence of direct health effects
from these pollutants, this report placed greater emphasis on
literature dealing with chemical and biological contaminants of
more direct public health concern.
85
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TABLE 16. LITERATURE REVIEWED PERTAINING
TO ACTIVATED SLUDGE
Contaminant Reference Number
Water Quality Parameters
Ammonia 4, 10, 59, 64, 113, 196, 277, 351, 352,
364, 385, 454, 614, 618, 799, 922, 1115,
1333, 1351
BOD 3, 10, 113, 146, 196, 209, 276, 352, 364,
560, 614, 623, 745, 799, 814, 864, 875,922,
1003, 1084, 1240, 1337, 1360, 1435, 1493
COD 196, 277, 364, 560, 601, 623, 668, 799,922,
1084, 1158, 1333, 1493
Chlorides 114, 819, 899, 1062
- Cyanides 623, 824, 922, 1062
Fluorides 1062
Nitrates 10, 59, 64, 113, 122, 352, 364, 375, 447,
454, 1351
Nitrites 10, 122, 352, 454, 1351, 1516
Oil and grease 810, 922, 1062, 1516
Phosphates 10, 13, 62, 63, 64, 113, 114, 196, 277,
352, 364, 375, 410, 454, 510, 513, 555,
601, 614, 763, 785, 814, 899, 922, 1213,
1230, 1337, 1341, 1349
Suspended solids 3, 10, 113, 146, 196, 227, 352, 560, 623,
668, 745, 922, 1084, 1435, 1516
Total dissolved 196, 276, 352, 500, 899
solids
Total organic 10, 113, 114, 352, 615, 1158, 1333, 1516
carbon
Other (general) 385, 391, 393, 531, 602, 622, 1310, 1333
Biological Contaminants
Aluminum 513, 814, 1004
86
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TABLE 16 (continued)
Contaminant Reference Number
Arsenic 1062
Barium 1062
Boron 352, 899, 1062
Cadmium 227, 228, 230, 922, 982, 992, 1004, 1062,
1063, 1323
Chromium 69, 71, 227, 228, 623, 922, 940, 969,
982, 1062, 1323
Cobalt 353, 982
Copper 69, 71, 227, 228, 230, 623, 875, 922,
982, 1004, 1062, 1063, 1323
Iron 227, 228, 353, 982, 1004, 1062, 1323
Lead 227, 228, 230, 922, 982, 1004, 1062,
1063, 1323
Manganese 227, 228, 410, 982, 1062, 1323
Mercury 227, 228, 471, 982, 992, 1004, 1062
Molybdenum 1323
Nickel 69, 71, 227, 230, 623, 922, 982, 1323
Seleni urn 1062
Zinc 69, 71, 227, 228, 922, 992, 1004, 1062,
1063
Other (general) 1062, 1323
Pesticides
Aldrin 353
DDT 1493
Synthetic/Organic 81fi 1(lcn
f\ i • J. *-J -> U » I "T _/ \J
Contaminants
87
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TABLE 16 (continued)
Contaminants Reference Number
Biological Contaminants
Bacteria 79, 161, 694, 1107
Cl ostrich'urn welchi 61 5
Coliforms 113, 568, 615, 983, 1084
Coxsackie virus 97, 161, 890
(A & B)
Escherichia col i 568
Fecal streptococci 568, 615, 1084
Mycobacterium 428, 568, 615
Parasitic worms «61, 428
Polio virus 97, 99, 161, 839, 840, 890
Protozoa 428
Salmonella 161, 197, 428, 568, 615, 717
Shigella 161 , 568, 615
Vibrio cholerae 161
Virus 79, 92, 95, 96, 100, 161, 352, 380, 382,
400, 428, 468, 492, 900, 1010, 1510
Other (general) 161, 531, 622, 1493
-------
The various activated sludge processes are all able to
remove over 90 percent of the soluble BOD found in wastewater.
Mitchell (922) recorded data on the removals achieved with
activated sludge processes of these and other water quality
contaminants during his study of the Los Angeles Hyperion
Treatment Plant as shown in Table 17.
TABLE 17 . ACTIVATED SLUDGE TREATMENT POLLUTANT
REMOVALS, LOS ANGELES, CA (922)
COD
BOD5
Suspended Solids
Oil and Grease
Phenol s
Ammonia nitrogen
Phosphorus
Cyanide
Treatment
Inf 1 uent
(mg/£)
315
165
103
28
0.09
20
10.1
0.30
Treatment
Effluent
(mg/£)
31
9
9
0.5
0.009
9.6
3.3
0.13
Percent
Removal
90
95
91
98
90
52
67
57
As can be seen from this table, relatively high removals of
most water quality contaminants can be attained in a practical
application over an extended period of time. These removal
efficiencies are supported by other researchers and reviewers,
including Noland and Birkbeck (1003), Huang et al. (602), Rickert
and Hunter (1158), Lindstedt and Bennett (799), and Besik (114).
ELEMENTAL CONTAMINANTS
Although the activated sludge process efficiently removes
biodegradable organic materials, only limited removal of soluble
elemental contaminants from the wastewater stream can be
achieved. The removal of elemental contaminants is governed by
two basic mechanisms: (1) the precipitation of metal hydroxides;
and (2) the adsorption of elemental contaminants by the activated
sludge floe. In either case, the elemental contaminants removed
will be contained in the sludge.
When suspended solids removals were in the 90 to 95 percent
range, 90 percent of the aluminum, iron, mercury, lead, and zinc
settled readily with the biofloc, according to Nomura and Young
(1004). Chromium (VI) and nickel median removals of 77 percent
and 50 percent, respectively, were recorded under the same
conditions. Morgan (969) stated that the removal percentages
for chromium could vary, depending upon the treatment process
used. Chromium exists in sewer systems in Cr (III) and Cr (VI);
concentrations depend upon pH. Cr (III) is readily adsorbed on
89
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particles or precipitated as Cr(OH)3(S). Chromium entering as
Cr (VI) experiences a strongly reducing environment (little or
no dissolved 0?) in sewers and treatment plants, and is thus
reduced to Cr (III) and either precipitated or adsorbed. Aera-
tion used in activated sludge or stabilization processes can
cause the resolubilization by oxidation of trivalent to hexava-
lent chromium, with resulting effluent water degradation.
The association of trace metals with suspended solids during
the activated sludge process was investigated by Chen and
Hendricks (227) and Chen and Lockwood (228) at the Hyperion
Their work confirms that many trace metals are
suspended solids, although the concentration of
particles does not appear to depend significantly
Rather, the removal of trace metals from the
Treatment Plant
associated with
trace metals on
on particle size
waste stream by
extent upon the
activated sludge processes depends to a great
adsorptive capability of the activated floes.
Mitchell (922) has recorded elemental removal efficiencies
for the Hyperion Plant over a three-year period, as shown in
Table 18 .
TABLE 18 . REMOVALS OF TRACE METALS
BY ACTIVATED SLUDGE PROCESSES,
LOS ANGELES, CA (922)
Element
Copper
Zinc
Silver
Nickel
Lead
Arsenic
Cadmium
Chromium
Inf1uent
(mg/l)
0.25
0.42
0.019
0.24
0.07
0.017
0.02
0.37
Effluent
(mg/l)
0.
0.
0.
0,
0,
0,
0,
08
23
012
15
08
013
013
0.013
Percent
Removal
68
46
37
38
24
35
96
These figures can be compared with the removal percentages as
shown in Table 19 (influent metal concentrations averaged less
than 1 mg/1). Clearly, the activated sludge process can reduce
but will not eliminate, trace metal concentrations in the muni-
cipal wastewater stream.
90
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TABLE 19 . PERCENT REMOVALS OF TRACE METALS BY
THE ACTIVATED SLUDGE PROCESS (1323)
Average
Element Percent Removal
Cadmium 56
Chromium 36
Copper 47
Iron 46
Lead 36
Manganese 16
Molybdenum 15
Nickel 20
Silver 48
Zinc 56
SYNTHETIC/ORGANIC CONTAMINANTS
Malaney et al . (836), in a study of the removal of possible
carcinogenic organic compounds by activated sludge, concluded
that no significant reduction was accomplished within normal
detention times at the three treatment plants studied. The
following Table 20 lists the possible carcinogens included in
the analysis.
Recent work by Wachinski et al. (1439) suggests that herbi-
cide detoxification can be achieved with a pure oxygen-activated
sludge treatment system that was determined to be both economical
and ecologically safe. A proprietary strain of mutant micro-
organisms, PHENOBAC (developed by the Worne Biochemical Corp.),
was utilized that was able to degrade halogenated phenols. Even
with relatively high herbicide concentrations (1380 mg/£),
degradation of as much as 73 percent was accomplished after a
16-day aeration period using optimum proportions of required
nutrients, microflora, and oxygen. According to the authors,
this figure represents a conservative estimate of possible
reductions, since testing was conducted at 18°C, while the
optimum growth temperature for PHENOBAC is close to 30°C.
BIOLOGICAL CONTAMINANTS
Activated sludge followed by secondary sedimentation can
remove over 90 percent of coliform or pathogenic bacteria that
remain after primary sedimentation; other biological pathogens
are removed to varying degrees. Nonetheless, even with 90
percent removal, appreciable amounts of pathogens remain present
in the effluent.
91
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TABLE 20. POSSIBLE CARCINOGENS
INCLUDED IN THE ANALYSIS (836)
2,3 - Butylene oxide
B - Propiolactone
Thiourea
Ethycarbamate
2 - Thiouracil
4 - Ethoxyphenylurea
B e n z i d i n e
4,4' - Dihydroxy-a,b-diethylsti1bene
2 - Naphthylamine
4,4' - Bis (dimethyl ami no) benzopheuone
p-Phenylazophenol
p-Phenylazoaniline
9,10 - Dimethylanthracene
1,2 - Benzanthracene
7 - Methyl-1,2-benzanthracene
9,10 - Dimethyl-1,2-benzathracene
1,2,5,6 - Dibenzanthracene
3/4 - Benzpyrene
1,2,4,5 - Dibenzpyrene
20 - Methylcholanthrene
2 - Nitrofluorene
2 - Fluoreneamino
N-2 - Fluorenylacetamide
7,9 - Dimethylbenz (c) acridine
7,10 - Dimethylbenz (c) acridine
Dibenz (a,h) acridine
Dibenz (a,j) acridine
92
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Pathogens can either be removed by adsorption onto the
sludge floes or destroyed by the predatory activity of the
zoogleal component. A review of the literature (531, 622)
reveals discussion of general wastewater removal rates with
little differentiation between removals and the biocidal prop-
erties of activated sludge.
Foster and Engelbrecht (428) provided a review of pathogen
removal by activated sludge processes. Conclusions of this
review are summarized in Table 21 .
TABLE 21 . REMOVAL OF PATHOGENS BY
THE ACTIVATED SLUDGE PROCESS (428)
Pathogen Percent Removal
Salmonella 96 to 99
Mycoba'cterium Slight to 87
Amoebic cysts No apparent removal
Helminth ova No apparent removal
Virus 76 to 99
Ova of intestinal parasites are apparently unaffected by
the activated sludge process; in fact, the literature indicates
that activated sludge-mixed liquor provides an excellent hatch-
ing medium for eggs.
A review by Hunter and Kotalik (615) provided additional
pathogen removal data, as summarized in Table 22 .
TABLE 22 . PERCENT REMOVALS OF BIOLOGICAL PATHOGENS BY
THE ACTIVATED SLUDGE PROCESS (615)
Pathogen Percent Removal
Coliform 90 to 99
Fecal streptococci 84 to 94
Shigella 90 to 99
Salmonella 70
Pseudomonas aergglnosa 99
Clostridlum perfringlrTs 90 to 99
MycobacteTTum tuberculosis 66 to 88
A 11st of the species of protozoans, nematodes, and fungi that
have been found 1n activated sludge effluent was also presented
by the authors. Species recorded Include:
93
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Protozoa -
Amoeba sp.
Euplotes patel1 a
Loxophyl1qm he! us
Oi komonas sp^
Pelodini urn reni forme
Phyllomitus anylophagus
Trlgonomonas compressar
Vortice11 a campanula
Epi sty!i s pii ciati1s
Nematodes -
Rhabditidae
Diplogasteridae
Dorylaimidae
Monochidae
Alternaria
Aspergillus
Aureobasi dium
Candida
Cryptococcus
F u s a r i u m
Geotrichum
Hansenula
Kloeckera
Mucor
P e n i c i 11 i u m
Rhodotorula
Saccharomyces
Torulopsis
Tri chodsrma
Trichosporon
The removal of viral
recently become the topic
viral removal of up to 90
activated sludge process.
contaminants by activated sludge has
of considerable research. In general
percent has been observed after the
However, large variations in removal
have been reported, probably because sampling was not temporally
coordinated (468). Destruction by sewage microflora and virus
adsorption to floe during the process are believed to be the
main factors governing viral removal. Table 23 reports typical
viral removals that can be expected from activated sludge
treatment (161).
TAB
Pat
LE 23 .
hogen
V
IRAL
REMOVAL
B
Y ACTIVATED S
Percent Remov
LU
al
DG
E
TREATMENT
(161)
Enteroviruses
Polio viruses
Coxsackie viruses
ECHO viruses
0 to 90 percent
0 to 90 percent
0 to 50 percent
no apparent removal
Malina et al. (839) concluded from their research that
viral inactivation by activated sludge is independent of the
hydraulic detention time. Polio virus adsorption to sludge is
almost immediate; the adsorbed virus particles are inactivated
according to first order kinetics with a half-life in the range
of 4 to 5 hr. Activated sludge utilizes aeration for optimum
dispersion of the flocculant masses which, along with gases
produced during microbial respiration, may entrain bacterial
and viral pathogens in aerosols. The aerosols released are in
the 5-ym diameter range, small enough to permit lung penetration
of a substantial proportion of the particles.
94
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SECONDARY TREATMENT: TRICKLING FILTER
INTRODUCTION
Trickling filters have been widely used for secondary bio-
logical treatment of municipal wastewater, and substantial
literature is available on this porcess, as indicated on Table
24 .
The trickling filter system generally consists of a tank
open on both top and bottom. The tank is filled with a rock or
plastic filter media having a high surface area to allow attach-
ment of zoogleal slimes and void fraction for movement and dif-
fusion of 'oxygen. Contaminant removal is accomplished through
adsorption at the surface of the biological slimes covering the
filter media. Following adsorption, the organics are utilized
by the slimes for growth and energy. The trickling filter is
followed by clarification to remove biological solids periodi-
cally flushed from the filter.
Trickling filter influent is usually from a primary treat-
ment system. System outputs include effluent, sludge, and pos-
sible aerosols.
The literature generally refers to percent removal, with
no distinction made between separation and degradation or des-
truction. As in the case of other secondary processes, the
literature primarily addresses the general and biological con-
taminants and is sparse in the areas of elemental, synthetic,
and biocidal contaminants.
WATER QUALITY PARAMETERS
The removal of BOD and suspended solids by trickling fil-
ters is reported to be from 65 to 95 percent, averaging about
85 percent (622, 1323). The efficiency of trickling filtration
decreases as temperatures fall below 2QOC. Imhoff et al . (622)
reported that a reduction of temperature from 20° to 10°C
results in an efficiency loss' of about 40 percent. Nickerson
et al. (996) found that chemical addition ahead of primary
clarifiers increases overall BOD and suspended solids removals
in trickling filters. Lager and Smith (745) reported that no
significant removal of total nitrogen or phosphorus occurred
during the conventional trickling filter process. In low-
rate filters, ammonia and nitrogenous organic compounds are
95
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TABLE 24. LITERATURE REVIEWED PERTAINING
TO TRICKLING FILTERS
Contaminant Reference Number
Water Quality Parameters
Ammonia 5, 41, 349, 745, 891, 1403, 1484
BOD 60, 241, 622, 745, 814, 907, 996, 1013,
1143, 1323, 1360, 1403, 1546
Chlorides 907, 1484
COD 241, 491, 891, 907, 1485
Nitrates 4, 5, 349, 745, 891, 1484
Nitrites 349, 891, 1484
Phosphates 70, 633, 745, 814, 891, 1013, 1484
Suspended solids 498, 622, 745, 996, 1013, 1323, 1403, 1546
Total organic 60, 241, 615
carbon
Other (general) 391, 393, 531, 622
Elemental Contaminants
Aluminum 814
Boron 1484
Cadmium 1323
Chromium 69, 1143, 1323
Copper 69, 1143, 1323
Germanium 1411
Iron 1323, 1484
Lead 1323, 1484
Manganese 1323, 1484
Molybdenum 1323
96
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TABLE 24 (continued)
Contaminant Reference Number
Nickel 69, 876, 1143, 1323
Zinc 69, 1143, 1323
Synthetic/Organic 241, 615
Contaminants
Biological Contaminants
Bacteria 79, 142, 161, 400, 486
Coliforms 486, 568, 983, 1278
Coxsackie virus 161
(A & B)
Escherichia coli 568
Fecal streptococci 1278
Mycobacterium 428
Parasitic worms 161, 428, 615
Polio vi rus 161
Protozoa 428
Salmonella 161, 428, 717, 1278
Shigella 161
Vibrio cholerae 161, 717
Virus 79, 95, 96, 97, 100, 161, 380, 429, 492,
1278
Other (general) 161, 486, 622
97
-------
usually oxidized to yield high proportions of nitrates and some
nitrites in the effluent; high-rate trickling filter effluents
are low in nitrates and nitrites due to limited system oxidation
Barth et al. (70) reported total phosphorous removals of up to
75 percent with direct dosing of aluminate to a trickling filter
ELEMENTAL CONTAMINANTS
Removal of elemental contaminants by trickling filters is
not well documented. A summary of the literature on metal
removals as compiled by the State of California Water Resources
Board is shown in Table 25.
TABLE 25. TRICKLING FILTER PROCESS REMOVAL
OF TRACE METAL CONTAMINANTS (1323)
Average
Element Percent Removal
Cadmium 5
Chromium 19
Copper 47
Iron 46
Lead 36
Manganese 16
Molybdenum 15
Nickel 20
Silver 48
Zinc 56
Trace metal removals by trickling filters are substantially
lower than those achieved with the activated sludge process be-
cause there is less formation and sedimentation of trace metal
complexes.
BIOLOGICAL CONTAMINANTS
Trickling filters do not effectively remove many biological
pathogens. Table 26 illustrates reported removal efficiencies.
Organisms are adsorbed into the zoogleal slime but due to
similar surface charges and morphology, biocidal effects are
variable (622).
98
-------
TABLE 26. REMOVAL OF PATHOGENS BY
TRICKLING FILTERS (428)
Pathogen Group
Efficiency
Salmonel 1 a
Mycobacterium
Amoebic Cysts
Helminth Ova
Virus
88
66
11
62
0
to
to
to
to
to
99
99
99
76
84
.9
.9
Foster and Engelbrecht (428) observed that trick-
ling filters are capable of reducing paratyphoid organisms by
84 to 99 percent. A review by Hunter and Kotalik showed 99.7
percent removal ofSchistosoma mansoni ova. The authors also
concluded that trickling filter effluents can contribute a
major portion of the free living nematode population found in
receiving waters.
Improperly operated low-rate trickling filters can provide
an excellent breeding area for insects, especially filter flies
(Psychoda) and springtails. Trickling filters cannot be depen-
ded upon to produce significant or consistent viral reductions.
Foster and Engelbrecht (428) reported removals ranging from 0
to 84 percent. Berg (97) speculated that even when viruses are
adsorbed, they may eventually be replaced by other substances
and leach out of the filter slime as a result of an equilibrium
effect.
99
-------
SECONDARY TREATMENT: AERATED LAGOONS
INTRODUCTION
Aerated lagoons are ponds in which mechanical aeration
is used to increase the rate at which oxygen is made avail-
able to facilitate biological stabilization. The aeration
also provides mixing for suspension of microbial floe.
The biological process does not include algae, and organic
stablization depends on the mixed liquor that develops within
the pond. Literature surveyed concerning aerated lagoons is
1isted in Table 27.
WATER QUALITY PARAMETERS
BOD removal by aerated lagoons is a function of aeration
period, temperature, and the nature of the wastewater. The
aeration of a typical domestic wastewater for five days at 20°C
provides about 85 percent BOD reduction; lowering the temperature
to 10°C reduces the efficiency to approximately 65 percent (531).
BIOLOGICAL CONTAMINANTS
A discussion of the literature by Parker (1052) revealed
that coliform reductions in the range of 80 to 99 percent can be
achieved with optimum detention time. This is supported by the
experiments of Carpenter et al. (198), who reported that coli-
form organisms are efficiently removed by the use of aerated
lagoons. Klock (717) stated that the coliform survival rate
in lagoons is a function of the oxidation-reduction potential
and temperature.
Berg (95) discussed the removal of viruses by stabilization
ponds, concluding that virus removal can be expected to be
erratic.
100
-------
TABLE 27. LITERATURE REVIEWED PERTAINING
TO AERATED LAGOONS
Contaminant Reference Number
Water Quality Parameters
Ammonia 454
BOD 91 , 484, 531, 651, 784, 1404
COD 91, 266, 484
Nitrates 454
Nitrites 454
Phosphates 328, 784
Suspended solids 484, 745
Other (general) 393, 531, 622
Biological Contaminants
Bacteria 198, 717
Coliforms 198, 568, 717, 1052
Escherichia coli 1052
Fecal streptococci 568, 1052
Salmonella 568
Virus 95, 198, 382
101
-------
SECONDARY TREATMENT: PONDING
INTRODUCTION
An oxidation or facultative pond is generally a shallow
earthen basin designed to promote a symbiotic existence between
algae and bacteria (745). Algal photosynthesis and surface
reaction maintain aerobic conditions in the photic region, while
anaerobic bacteria flourish in the aphotic zone. Ponds are
normally operated in series and are sometimes used for
"polishing" effluent from conventional secondary processes.
Influent to a ponding system may be raw sanitary waste,
primary effluent, or secondary effluent. Pond effluent can
enter the environment by direct discharge or by seepage to the
groundwater. Literature reviewed concerning ponding is listed
in Table 28. As can be seen from this review, only a limited
amount of research has examined the removal of contaminants by
the ponding process.
WATER QUALITY PARAMETERS
Oxidation pond removal efficiencies for suspended solids
and BOD can vary widely and may even reach negative values
(745). Removal ranges of 60 to 50 percent have been reported
for suspended solids, and of 70 to 10 percent for BODg. This
variation occurs because most influent BOD is converted into
suspended algal mass. This mass exerts a BOD demand and provides
suspended solids that may be carried into the effluent.
Bacterial decomposition and algal growth are both retarded
by reduced temperatures, reducing removal efficiency of the
ponding process (531).
SYNTHETIC/ORGANIC CONTAMINANTS
The removal of trisodium nitri1otriacetate (NTA) by ponding
has been investigated by Klein (714). He found that after a
two-month acclimation period, steady-state removal was in excess
of 90 percent, with influent concentrations in the range of
30 mg/£.
102
-------
TABLE 28. LITERATURE REVIEWED PERTAINING TO PONDING
Contaminant Reference Number
Water Quality Parameters
BOD 326, 531, 745
COD 1484
Chlorides 427
Fluorides 427
Nitrates 226, 326, 375, 446, 1432
Phosphates 326, 375, 426, 446, 1484, 1500
Suspended solids 393, 745
Other (general) 393, 531, 745
Synthetic/Organic 714
Contami nants
Biological Contaminants
Bacteria 842, 1278
Coliforms 849, 850, 1278
Escherichia coli 850, 1048
Fecal streptococci 849, 850, 1278
Salmonel1 a 669
Virus 95, 96, 97, 1278
103
-------
BIOLOGICAL CONTAMINANTS
Ponding preceded by adequate secondary treatment provides
excellent bacteria removal. Kampelmacher and Jansen (669) found
that removal of salmonella by oxidation ponds was not Inferior
to removal achieved by conventional treatment plants. Species
of the coliform group, although reduced by ponding, are not
effectively eliminated according to Parhad and Rao (1048).
Slanetz et al. (1278), however, reported that if two ponds were
operated in a series at temperatures of 17° to 26 C, the die-off
rate of coliform, fecal coliforms, and fecal streptococci ranged
from 95 to 99 percent. During winter when temperatures were in
the 1° to IOC range, the die-off rate was 46 times lower. Berg
(97) states that virus removal by ponding is erratic, ranging
from 0 to 96 percent; virus recovery decreased as the effluent
passed through a series of maturation lagoons.
104
-------
TERTIARY TREATMENT: CHEMICAL TREATMENT
INTRODUCTION
The purpose of chemical treatment is to coagulate sus-
pended solids and precipitate phosphate and various trace
metals. Chemical coagulation of secondary effluents may
be accomplished by the addition of lime, alum, polymers, or
iron salts, and involves three operations: (1) injection
and rapid mixing of the coagulants to neutralize the pre-
dominantly negative charges on suspended matter; (2) gentle
stirring to promote agglomeration of the coagulated particles
into large, settleable floe; and (3) sedimentation to provide
gravity separation of the flocculated material from the waste
water. The settled material is transferred to a sludge hand-
ling system. As indicated by Table 29, a great deal of infor-
mation is available concerning the removal of various public
health impairing contaminants by chemical treatment processes.
WATER QUALITY PARAMETERS
Culp and Shuckrow (278) investigated chemical treatment of
raw wastewater with lime and found that removals of 95 to 98
percent phosphorus and 99 percent suspended solids can be
achieved with chemical clarification followed by carbon adsorp-
tion. The treatment of municipal wastewater with alum precipi-
tation as studied by Shuckrow et al. (1254) resulted in
removal efficiencies of 85 percent for COD and 83 percent for
total organic carbon.
The removal of BOD, suspended solids, and phosphorus as
reviewed by Lager and Smith (745) is summarized in Table 30 .
Removals from secondary effluents of the magnitudes listed
obviously provide a high quality effluent.
105
-------
TABLE 29 . LITERATURE REVIEWED PERTAINING TO CHEMICAL TREATMENT
Contaminant Reference Number
Water Quality Parameters
Ammonia 43, 120, 277, 284, 390, 601, 614, 885,
891, 1321 , 1510
BOD 119, 124, 277, 278, 284, 334, 498, 614,
745, 794, 799, 813, 814, 878, 885, 996,
1012, 1013, 1215, 1337, 1388, 1393, 1402,
1467, 1510
Chlorides 885
COD 119, 120, 124, 249, 277, 278, 284, 329,
498, 885, 1254
Cyanides 107, 922, 993
Fluorides 410, 1321
Nitrates 119, 120, 124, 249, 277, 278, 284, 329,
498, 885
Nitrites 284, 390, 891, 1085, 1321, 1510
Phosphates 278, 284, 290, 295, 326, 328, 375, 410,
423, 479, 513, 601, 614, 650, 711, 743,
745, 763, 794,,799, 814, 865, 876, 885,
891, 1012, 1013, 1036, 1039, 1213, 1215,
1265, 1311, 1321, 1337, 1341, 1393, 1402,
1466, 1500, 1510, 1547, 1552
Suspended solids 38, 108, 119, 120, 249, 278, 330, 498,
601, 701, 745, 799, 813, 896, 996, 1012,
1013, 1215, 1254, 1338, 1389, 1393, 1402,
1485, 1510
Total dissolved 277, 1215, 1321
sol ids
Total organic 119, 120, 124, 284, 1254, 1467
carbon
Other (general) 385, 391, 393, 531, 622
106
-------
TABLE 29 (continued)
Contami nant Reference Number
Elemental Contaminants
Aluminum 334, 814, 1036, 1383
Antimony 38
Arsenic 38, 855, 922, 1062, 1274, 1323, 1383
Barium 38, 855, 1062, 1323, 1383
Boron 1383
Cadmium 38, 39, 107, 802, 855, 922, 1103, 1323,
1383, 1504, 1505, 1506
Chromium 38, 39, 204, 802, 855, 1323, 1385, 1504,
1505, 1506
Cobalt 1383
Copper 38, 39, 107, 204, 855, 922, 1323, 1383,
1504, 1505, 1506
Iron 38, 334, 763, 1323, 1383, 1504, 1506
Lead 855, 922, 1323, 1383, 1411, 1504, 1506
Manganese 38, 334, 410, 855, 1323, 1383, 1504,
1505, 1506
Mercury 38, 775, 811, 855, 922, 1077, 1323,
1353, 1383, 1504, 1505, i506
Molybdenum 38, 1323
Nickel 38, 855, 922, 1323, 1383, 1504, 1505
Selenium 802, 1323, 1383
Uranium 38, 1323
Zinc 38, 107, 855, 922, 1323, 1383, 1504,
1505, 1506
Other (general) 38, 760, 799, 1323, 1506
107
-------
TABLE 29 (continued)
Contaminant Reference Number
Synthetic/Organic 6, 1496
Contaminants
Biological Contaminants
Adeno virus 1522
Bacteria 79, 331, 335, 670, 799, 830, 891
Coliforms 101, 284, 469, 799, 896, 1393
Coxsackie virus 97, 1522
(A & B)
ECHO virus 1522
Escherichia coli 880, 1183, 1382
Fecal streptococci 799
Hepatitis virus 469, 1522
Parasitic worms 830
Polio virus 103, 159, 223, 499, 837, 1293, 1313,
1382, 1510, 1522
Protozoa 830
Salmonella 670
Virus 79, 92, 96, 97, 159, 219, 220, 245, 246,
259, 283, 284, 331, 380, 381, 468, 848,
1009, 1057, 1071, 1243, 1302, 1312, 1313,
1510, 1511
108
-------
TABLE 30. REMOVALS ACHIEVED BY CHEMICAL
CLARIFICATION (745)
Chemical BOD Suspended Solids
Removal Removal
(Percent) (Percent)
Lime pH 11.5 80 90
Ferric Chloride 80 95
170 mg/£ dose
Ferric Chloride 75
80-100 mq/l dose
Phosphate removal by chemical precipitation (lime) has
received considerable attention in the literature in recent
years. The research on phosphate removal in general indicates
that lime clarification usually provides removal efficiencies
greater than 90 percent. This is supported by the work of
Davis (295), Sturm and Hatch (1341), Johnson (650), and
Bernhardt et al. (108) .
ELEMENTAL CONTAMINANTS
The precipitation of metal hydroxides from solution is
governed by the pH and the concentration of the metal ion in
solution. Since many of the trace metals form insoluble hydro-
xides near pH 11, lime coagulation results in a reduction of
these metal concentrations. Table 31 (1323) summarizes the
effects of lime coagulation on a number of heavy metals. Some
of the data were collected on industrial .metal wastes charac-
terized by metal ion concentrations a great deal higher than
would occur in any municipal plant influent. The data are
included here as examples of possible metal reductions, since
since such figures from chemical coagulation of municipal waste
waters are scarce.
As can be seen from these figures, arsenic, molybdenum,
and selenium had relatively poor removal rates,and the poten-
tial removal of mercury was estimated to be low. Only 11 per-
cent of hexavalent chromium was removed, although the trivalent
form was reduced more than 99.9 percent. Most other metals
tested were very effectively reduced at high pH. Lower remo-
vals of these same metals (usually less than 50 percent) can be
achieved with alum coagulation at near neutral pH values
(1323), a fact that illustrates the dependence of precipitation
on pH.
Gulledge and O'Connor (520) studied the removal of arsenic
(V) from tap water in jar tests using alum and ferric sulfate.
109
-------
TABLE 31 . REMOVAL OF ELEMENTAL CONTAMINANTS
BY LIME COAGULATION (1323)
Metal
Antimony*
Arsenic3
Barium9
Bi smuth3
Cadmium
Chromium (+6)
Chromium (+3)
Copper
Gold9
Iron
Leada
Manganese
Mercury3
Molybdenum
Concentration
Before Treatment
mg/i
23
--
Trace
0.0137
0.56
7,400
15
15,700
7
7
302
15
--
13
17
2.0
15
2.3
2.0
21.0
Trace
11
Concentration
After Treatment
mg/£
--
23
1.3 (sol)b
.0002 (sol)
0.00075
0.050
2.7
0.4
0.79
1
.05
Trace
0.6
<.001 (sol)
2.4
0.1
1.2C
<.001 (sol)b
0.5
<0.1
l.ic
0.05
Oxide Soluble
9
Percent
Removal
90
<10
0
Abt. 50
94.5
11
99.9 +
97
99.9 +
86
93
99 +
97
90 +
82
99 +
40
90+
97
96
45
95
<10
Abt. 10
18
110
-------
TABLE 31 (continued)
Metal
Nickel
Selenium
Silver
Teluri uma »d
Titanium3 >d
Urani ume
Zinc
Concentration
Before Treatment
mg/£
160
5
5
100
16
0.0123
0.0546
--
--
17
Concentration
After Treatment
mg/l
0.08
0.5
0.5
1.5
1.4
0.0103
0.0164
(<0.001?)
(<0.001?)
?
0.3
.007 (sol)
% Removal
99.9 +
90
90
99
91
16.2
97
( 90+)
( 90+)
?
98
90+
a The potential removal of these metals was estimated from
solubility data.
b Barium and lead reductions and solubilities are based
upon the carbonate.
These data were from experiments using iron and manganese
. in the organic form.
Titanium and telurium solubility and stability data made
the potential reduction estimate unsure.
Uranium forms complexes with carbonate ion. Quantitative
data were unavaiTable to allow determination of this
effect.
Ill
-------
In each case, increased coagulant dosage resulted in increased
arsenate removal. Dependence on pH was again seen in this
study. Under optimum conditions, removals of 85.8 percent with
alum and 96.8 percent with ferrous sulfate were achieved.
Pilot plant studies of municipal wastewater containing 5
mg/l arsenic cited by Patterson (1062) suggest that chemical
treatment can provide efficient removal of this element.
Ferric sulfate at 45 mg/£ Fe and pH 6.0 removed 90 percent of
the arsenic; lime at 600 mg/£ and pH 11.5 removed 73 percent.
In similar studies cited by Patterson, barium removals of 97
percent were obtained when municipal wastewater dosed with 5
mg/£ barium was treated with 45 mg/£ Fe at pH 6.0. Lime at
600 mg/£ and pH 11.5 resulted in 80 percent removal. The re-
moval of cadmium from waters by sorption onto hydrous oxides of
solid metals such as manganese and iron was investigated by
Posselt and Weber (1103). They concluded that sorptive uptake
of cadmium on such materials would constitute a method easily
adaptable to present treatment technology.
BIOLOGICAL CONTAMINANTS
Chemical treatment can be used to reduce or remove many
biological pathogens present in municipal wastewater. Lindstedt
and Bennett (799) evaluated the effectiveness of lime clarifi-
cation in reducing bacterial concentrations, finding that
treatment effectiveness increases with increasing chemical
dosage and pH. At a lime dosage of 400 mg/l , fecal coliform,
fecal streptococcus, and total coliform concentrations could be
reduced by two orders of magnitude. It was also found that
about 90 percent removal of bacteria can be achieved through
alum clarification over a broad range of alum dosage.
Jar tests employing the f2 bacteriophage virus, lake
water, and a variety of chemical coagulants and polyelectrolyte
coagulant aids were conducted by York and Drewry (1522). As
shown in Table 32 , aluminum sulfate (alum), ferric chloride,
ferric sulfate, ferrous sulfate, and polyelectrolyte B were
found to give maximum virus removals greater than 90 percent at
optimum dosage.
112
-------
TABLE 32 . COMPARISON OF THE EFFECTIVENESS OF
THE COAGULANTS TESTED (1522)
Coagulants-
Coagulant Aids
A12(S04)3
FeCl3
Fe2(S04)3 x H20
FeS04 and Ca(OH)2
Al 2( 50^)3 and
N52OAT 2U3
Polyelectrolyte A
Al2(S04)3 and
polyelectrolyte B
Polyelectrolyte B
A12(S04)3 and
polyelectrolyte B
Al2(S04)3 and
polyelectrolyte C
A12(S04)3 and
polyelectrolyte E
A12(S04)3 and
polyelectrolyte F
^12(504)3 and
polyelectrolyte D
Dose
mg/t
25
50
50
36
30
23
2.0
18
1.0
2.0
18
0.7
18
0.4
18
0.1
18
0.1
18
1.0
Maximum
Virus
Removal
percent
99.9
99.4
92.0
93.5
98.6
76
99.2
99.6
99.8
99.3
99.3
99.6
99.4
Berg et al. (103) experimentally mixed polio virus I and secon-
dary effluent in containers, added lime, and stirred the solu-
tion for 15 min to allow formation of floe particles. Settling
to 75 min. The removals obtained with varying
are shown in Table 33 . Large coagulant doses
effecting virus removals of up to 99.9 percent.
After further investigation, the authors concluded that signi-
ficant destruction of viruses can be attributed to the high
pH occurring with high lime concentrations (in the range of 400
to 500 mg/t).
followed for 60
dosages of lime
were capable of
113
-------
TABLE 33 . REMOVAL OF POLIO VIRUS I FROM SECONDARY
EFFLUENT BY FLOCCULATION WITH Ca(OH)2 (103)
Ca(OH)2
Concen-
tration
mg/£
200
300
400
500
500
Initial Virus
Concen-
tration
pfu/£
33,333
51,480
55,000
33,333
33,333
pH of
Treated
Effluent
9.30
10.21
11.30
11.03
11.01
Surviving Virus
Concen-
tration
pfu/£
2,200
15,912
1,940
505
47
Virus
Removal
percent
92.3
69.1
96.5
98.5
99.86
Chaudhuri and Engelbrecht (219) used water devoid of
extraneous organic matter in their laboratory investigation of
virus removal with alum. Using the experimentally determined
optimum pH and dosage, 98.0 percent removal of bacteriophage
T4 and 99.9 percent removal of bacteriophage MS2 were obtain-
able. However, the addition of organic matter in the form of
albumin or treated wastewater lowered these efficiencies. For
example, only 95.7 percent removal of bacteriophage T4 was ob-
tained after the addition of 200 ml/I of settled wastewater.
Further experiments demonstrated no inactivation of the virus
particles that were removed in the settled floe. However,
Brunner and Sproul (159) demonstrated 60 percent permanent Inac-
tivation in the case of polio virus I removed from solution
with aluminum phosphate precipitates. Their studies of polio
virus I and bacteriophage T2 removal showed that under optimum
conditions, removals can reach 98 and 94 percent, respectively,
with aluminum and calcium precipitation. Actual removals are
related to pH and chemical dosage.
Chemical treatment (high pH) holds considerable promise
as a means of effectively inactivating or destroying pathogenic
organisms contained in wastewater. By itself, chemical treat-
ment cannot be relied upon to produce a pathogen-free effluent;
used in conjunction with disinfection, however, it can help
ensure that such an effluent is achieved.
114
-------
TERTIARY TREATMENT: FILTRATION
INTRODUCTION
Inability of gravity sedimentation in secondary clarifiers
to remove small particles (and associated public health impair-
ing contaminants) is a limitation of BOD and suspended solids
removal by conventional wastewater treatment. Filtration as a
tertiary process upgrades treatment performance by removing a
portion of the unsettled suspended solids from secondary efflu-
ents. In addition, filtration often precedes other tertiary
processes such as adsorption and ion exchange since the pre-
sence of suspended solids interferes with the operation of
these processes. Reviewed literature on filtration is shown in
Table 34.
Filtration of wastewater to reduce the suspended solids
concentration is accomplished by passage through a bed of gran-
ular particles. Single, dual, or mixed media beds may be used,
composed of anthracite coal, granite, sand, and/or gravel
(531). Suspended solids are removed by a variety of mecha-
nisms: straining, impingement, settling, and adhesion. The
treatment efficiency of the process is influenced by:
• The concentration and characteristics of the wastewater
solids (particle-size distribution, surface character-
istics, organic versus inorganic, etc.)
• The characteristics of the filter media and filtering
aids used (particle-size distribution, surface charac-
teristics, etc.)
• The design and operation of the filter.
Since wastewater flow rate and solids content are variable, and
processes upstream of filtration may vary in performance, the
efficiency of filtration may also be expected to vary. For
this reason, values presented in the following discussion
should be considered as merely indicative of the range of
achievable removals.
WATER QUALITY PARAMETERS
In general, the best effluent quality achievable by
plain filtration of secondary effluent is about 5 to 10 mg/fc
for suspended solids and BOD. The suspended solids content of
115
-------
TABLE 34. LITERATURE REVIEWED PERTAINING TO FILTRATION
Contaminant Reference Number
Water Quality Parameters
Ammonia 43, 548, 885, 1323, 1546
BOD 119, 314, 697, 771, 885, 1319,1323,
1393, 1546
COD 119, 771, 885, 1319, 1323
Chlorides 114, 885, 1062, 1319, 1367
Cyanides 1062
Fluorides 1062
Nitrates 384, 885, 1319, 1367
Oil and grease 1062
Phosphates 108, 119, 314, 659, 885, 1323, 1393
Suspended solids 39, 108, 454, 531, 659, 771, 1323, 1367,
1393, 1452, 1546
Total dissolved 1062
solids
Total organic 114, 119, 1323
carbon
Other (general) 385, 531, 622
Elemental Contaminants
Arsenic 1062, 1244
Barium 1062
Boron 1062
Cadmium 38, 39, 802, 1062, 1063, 1559
Chromium 38, 39, 802, 1062, 1559
Copper 38, 39, 697, 1062, 1063, 1319, 1559
Iron 38. 1062. 1319, 1559
116
-------
TABLE 34 (continued)
Contaminant Reference Number
Lead 1062, 1063, 1559
Manganese 38, 635, 1062, 1319, 1559
Mercury 1062
Nickel 38, 1062, 1063
Selenium 38, 802, 1062
Zinc 38, 697, 1062, 1063, 1319, 1559
Other 38, 1062
Synthetic/Organic 1062, 1496
Contaminants
Biological Contaminants
Bacteria 161, 331
Coliforms 101, 469, 608, 617
Coxsackie virus 95, 161
(A & B)
ECHO virus 1182
Hepatitis virus 469
Parasitic worms 161, 1469
Polio virus 103, 153, 161, 223, 1168, 1182, 1293
Salmonella 161
Shigella 161
Vibrio cholerae 161, 424
Virus 25, 95, 96, 152, 154, 161, 261, 331,
454, 468, 492, 608, 1009, 1240, 1312
Other (general) 153, 161
117
-------
secondary effluent was reduced to 5 mg/s, with both rapid sand
and mixed media filters, employed respectively at a treatment
and a pilot plant. Complete removal, however, could not be
effected. If further reduction is desired, chemical coagula-
tion must precede filtration (454, 531).
Two studies at a pilot facility and a major treatment
plant (1323) indicated that essentially complete suspended
solids removal was accomplished when filtration was preceded
by chemical treatment of secondary effluent. Chemical clar-
ifier effluent contained 0.7 mg/n total phosphorus; after
filtration, this phosphorus content was reduced to 0.1 mg/fc.
At the pilot facility, the filter effluent contained 17.6 mg/£
COD, 9 mg/i total organic carbon (TOC) , and no detectable
phosphorus, compared with filter influent concentrations of
18.1 mg/£ COD, 8.6 mg/s, TOC, and 0.4 mg/i phosphorus.
When chemically treated secondary effluent was applied to
rapid-sand filters, generally 20 to 60 percent of applied BOD
was removed, 30 to 70 percent of phosphate, and 40 to 80 per-
cent of the suspended solids (1393). These values were ob-
tained using a variety of influent concentrations and chemical
dosages. The results of 30 days continuous operation of a
pilot plant practicing filtration preceded by chemical treat-
ment of secondary effluent are summarized in Table 35 (1323).
TABLE 35. RESULTS OF ONTARIO, CANADA, PILOT PLANT
STUDY INVOLVING FILTRATION PRECEDED BY
CHEMICAL TREATMENT OF SECONDARY EFFLUENT (1323)
Qua!i ty Parameter
Raw
Wastewater
Secondary
Effluent
Filter
Effluent
Total Organic Carbon
110-165
14-28
4.5-7.5
BOD5 (mg/l)
P04 (as P04)
Total Nitrogen N
Ammonia N
Suspended Solids
230-400
9-21
27-51
17-29
148-268
5-14
1.3-3.5
25-37
21-29
13-37
2.0-3.0
0.4-1.0
20-35
18-29
3-12
Results of a pilot, plant study at Cleveland, Ohio - where
chemical coagulation and settling of raw wastewater was
118
-------
followed by filtration and granular carbon adsorption - were
reviewed by Gulp and Shuckrow (278). On the basis of the data
provided, removals attributable to chemical treatment and fil-
tration can be calculated to be 66 percent of the applied BOD
and 77 percent of the applied COD. A second pilot plant study
reviewed by the authors involved the same treatment scheme.
Removals at this plant due to combined chemical treatment and
filtration can be inferred to range from 77 to 84 percent.
While studying soil filtration, de Vries (3K) applied
primary effluent to a sand filter and obtained BOD and phos-
phate removals of nearly 100 percent. Phosphate removals were
attri buted
grains
100
to the natural coatings of
Fe203
and
A1203
-on sand
ELEMENTAL CONTAMINANTS
After chemical treatment, filtration to remove residual
particulate matter may provide some additional removal of ele-
mental contaminants. Elemental contaminant removals achieved
by filtration depend primarily upon the extent of suspended
solids removals, with which the various trace elements are
associated. Table 36 , compiled from the literature by Argo
and Gulp (38) gives results for sand filtration of some
municipal and industrial wastes.
TABLE 36 . HEAVY METAL REMOVAL BY SAND FILTRATION
FOLLOWING LIME COAGULATION (38)
Metal
Cd
Cr+6
Cr+3
Cu
Fe
Mn
Ni
Se
Concentration
Before Filt.
0
0
2
0
0
Trace to
.00075 mg/l
.0503
.7
.79
-
_
_
.08
.0103
Concentrati on
After Treat. pH
0
0
0
0
0
1
0
1
0
0
0
.00070
.049
.63
.32
.5
.1
.2 Organic
.1
.1 Organic
.1
.5
.00932
8.
7.
7.
8.
9.
10.
10.
10.
10.
8.
9.
11
1
6
6
7
5
8
5
8
5
7
5
% Removal
By Filt.
95
6
2
77
59
9
.6
.6
.5
.5
119
-------
TABLE 36 (continued)
Metal
Ag
Zn
Concentration
Before Hit.
0.00164
0.97
Concentration
After Treat.
0.00145
0.23
2.5
£H
11
8.7
9.5
% Remova
By Filt.
11.6
76.3
1
Patterson (1062) cites evidence from pilot plant studies that
little or no additional removal of arsenic was afforded by fil-
tration of chemically treated municipal wastewater.
BIOLOGICAL CONTAMINANTS
Several investigations have been reported concerning the
removal of viruses by sand filtration. It has been shown that
insignificant virus removal is achieved by rapid filtration
through clean sand (25, 95). However, virus removal efficiency
will be increased by impregnation of the filter medium with
coagulated floe, the presence of organic matter trapped in the
sand, chemical flocculation prior to filtration, or a decrease
in the filtration rate. The addition of iron salts prior to
filtration has resulted in significantly higher coliform reduc-
tions, as discussed by Hunter et al. (617). Similarly,
Robeck et al. (1168) noted that if a low dose of alum was fed to
a rapid coal and sand filter just ahead of filtration, more
than 98 percent of polio virus Type I could be removed. If the
dosage was increased and conventional flocculators and settling
were used, removal was increased to over 99 percent. The
authors also noted a general trend toward better removal of
polio virus I with slower filtration rates although their data
were often erratic. At slow sand filter rates (0.6 to 1.2
£/min/m2), removal ranged from 50 percent to about 98 percent.
At rapid filtration rates (38 to 76 £/min/m2), virus removals
ranged from about 10 to 70 percent. Similarly, research exam-
ined by Berg (95) showed that filtration through sand at 7.5 I/
min/m2 removed 99 percent of the coxsackie virus A5 while fil-
tration at 75 £/min/m2 removed only about 10 percent of the
virus .
Brown et al. (153) reported 70 to 90 percent removals of
low concentrations of either bacteriophage T2 or polio virus
Type I by filtration through uncoated diatomaceous earth.
However, no significant virus removals by uncoated diatomaceous
earth were achieved in a laboratory study by Amirhor and
Engelbrecht (25). With polyelectrolyte-coated filter media,
removals greater than 99 percent were consistently achieved (25)
120
-------
In laboratory tests by Berg et al . (103), from 82 to
greater than 99.8 percent of polio virus I was removed from
chemically treated effluent by rapid sand filtration. The
results of these tests are given in Table 37 .
TABLE 37 . REMOVAL OF POLIO VIRUS I FROM
Ca(OH)2 FLOCCULATED EFFLUENT BY RAPID
SAND FILTRATION
AS MEASURED BY MEMBRANE FILTER RECOVERY OF VIRUS
(103)
Test No.
Virus Concentration pfu/£
Before Sand
Fi1tration t
After Sand
Fi1tration
Virus Removal
percent
1
2
3
4
5
2,200
15,912
1,940
505
47
397
750
<4
12
2.8
82.0
95.3
>99.8
97.5
94.0
* Filtration rate 2.25 gpm/sq ft through 8 in of sand.
t Virus concentration in flocculated effluent just prior to
sand filtration.
Laboratory experiments on the removal of nematodes by
rapid sand filtration were conducted by Wei et al . (1469).
Removal efficiency was about 96 percent when all the nematodes
in the influent were dead or nonmotile. However, most motile
nematodes were able to penetrate the filter bed.
Sand
cysts and
Bryan (161)
ded.
filtration may also provide some removal of amoebic
ascaris eggs, according to a literature survey by
He did not indicate the levels of removal affor-
121
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TERTIARY TREATMENT: ADSORPTION
INTRODUCTION
Adsorption refers to the removal from water or wastewater
streams of dissolved contaminants by their attraction to and
accumulation on the surface of an adsorbent substance. Activa-
ted carbon is the most widely used adsorbent in municipal
wastewater treatment to remove trace organics. Adsorption
using activated carbon is utilized as a tertiary treatment
step, usually following sand or multimedia filtration.
Carbon adsorption systems generally utilize granular or
powdered activated carbon packed in a column or forming a fil-
ter bed through which wastewater is passed. Three consecutive
steps occur in the adsorption of wastewater contaminants by
activated carbon: (1) the film diffusion phenomenon, or the
transport of the adsorbate through a surface film surrounding
the activated carbon; (2) the diffusion of the adsorbate within
the pores of the activated carbon; and (3) adsorption on the
interior surfaces of the activated carbon.
Carbon adsorption of contaminants has been the topic of
many research projects as shown in the literature surveyed for
this report, tabulated in Table 38.
WATER QUALITY PARAMETERS
Adsorption is most effective for removing refractory and
other organics from wastewater. This is especially important
when effluents of exceptional quality are required (e.g., for
groundwater recharge or other reuse applications). Adsorption
can be used either as a polishing step or as the major treat-
ment process (1467). Rizzo and Schade (1165) and Zanitsch and
Morand (1554) reported that carbon columns alone were capable
of about 85 percent removal of BOD from wastewaters entering
the columns. Bishop et al. (119, 123) and Zanitsch and Morand
(1554) reported 75 to 80 percent TOC removals under the same
conditions. Weber et al . (1467) found that a treatment system
composed of primary settling, ferric chloride coagulation, and
carbon adsorption could remove up to 97 percent of the influent
BOD.
There is some disagreement among researchers regarding the
ability of activated carbon to remove nitrogen species from
122
-------
TABLE 38. LITERATURE REVIEWED PERTAINING TO ADSORPTION
Contami nant Reference Number
Water Quality Parameters
Ammonia 113, 119, 120, 196, 885
BOD 113, 119, 120, 156, 170, 196, 278, 315,
326, 498, 522, 745, 885, 1165, 1467, 1554
COD 119, 120, 156, 196, 278, 329, 498, 668,
885, 1165, 1547
Chlorides 114, 885, 1026, 1289
Nitrates 113, 119, 120, 326, 1467
Oil and grease 480
Phosphates 24, 113, 114, 119, 120, 196, 326, 448,
885, 1467
Suspended solids 39, 113, 119, 120, 170, 196, 278, 326,
498, 668, 745, 1165, 1442
Total dissolved 196, 454, 1547
solids
Total organic 113, 114, 119, 120, 123, 315, 522, 1086,
carbon 1165, 1262, 1467S 1554
Other (general) 531, 622
Elemental Contaminants
Aluminum 1383
Arsenic 855, 1383
Barium 855, 1383
Boron 1140, 1383
Cadmium 38, 39, 619, 753, 760, 855, 1383
Chromium 38, 39, 603, 619, 753, 760, 855, 1169,
1383
Cobalt 753, 1383
123
-------
TABLE 38 (continued)
Contaminant Reference Number
Copper 39, 620, 855, 1383
Iron 620, 753, 1383
Lead 753, 855, 1383
Manganese 753, 855, 1873
Mercury 753, 811, 855
Nickel 753, 855, 1383
Selenium 38, 1062, 1383
Zinc 620, 855, 1383
Other (general) 38, 760
Biocidal Contaminants
Chlorinated 762, 1221
hydrocarbons
Dieldrin 523
Herbicides 514, 516, 1221
Other (general) 523, 628, 762, 1323
Synthetic/Organic 123, 315, 454, 480, 523, 628, 651, 762,
Contaminants 862, 1153, 1290, 1323, 1465, 1496
Biological Contaminants
Bacteria 283
Coliforms 113
Escherichia coli 263
Polio virus 223, 466, 467, 1293, 1313
Virus 246, 263, 468, 492, 1010, 1312, 1313
124
-------
wastewaters. Bishop et al . (119) reported that carbon adsorp-
tion had little effect on nitrogen concentrations. On the
other hand, Weber et al. (1467) found that their primary set-
tling ferric chloride coagulation, carbon adsorption system re-
moved 95 percent of the influent nitrate; the reduction was
attributed partly to biological populations growing in the col-
umns .
Weber et al. (1467) reported that their pilot system also
removed 90 percent of the phosphate in the influent wastewater.
However, much of the research on phosphate adsorption has been
concerned with adsorbents other than carbon. Gangoli and
Thodos (448) reported that both Fl Alumina and fly ash were
capable of removing up to 99 percent of influent phosphate
levels. Ames and Dean (24) demonstrated that an alumina column
could treat up to 400 column volumes before reaching the 10
percent phosphorus breakthrough level.
Since color is contributed largely by organic compounds in
water, high color removal levels with adsorption should be pos-
sible. Zanitsch and Morand (1554) demonstrated 90 percent
color removal. They also noted an 86 percent suspended solids
removal, presumably by filtration.
ELEMENTAL CONTAMINANTS
Not a great deal of literature is available on the removal
of elemental contaminants by carbon adsorption. Such systems
are not specifically designed to remove ionic elemental con-
taminants, but some elementals are incidentally removed. When
the metallic contaminants are in an organometal1ic complex,
carbon adsorption columns can remove specific species. Litera-
ture from several sources (619, 760) reveals that high removals
(95 percent) of cadmium and hexavalent chromium by carbon ad-
sorption are possible. Huang and Wu (603) found that the effi-
ciency of chromium removal increases with decreasing chromium
concentration. While the mechanism of removal is not well
understood, Roersma et al. (1169) were able to describe the
reduction of Cr+6 to Cr+3. The Cr+6 is adsorbed within the
pores of the activated carbon which, in turn, is slowly oxi-
dized to C02> reducing the chromium ion.
Activated carbon treatment of a secondary-treated munici-
pal wastewater was found to reduce selenium from 9.32 to 5.85
ppb in a study cited by Patterson (1062). This represents a 37
percent removal efficiency. Patterson cited a second study in
which the selenium removal efficiencies of several advanced
wastewater treatment processes were determined in the treatment
of secondary effluent containing 2.3 ppb selenium. While sand
filtration alone reduced the effluent concentration by 9.5 per-
cent, sand filtration followed by activated carbon treatment
125
-------
provided a cumulative removal of 43.2 percent of the selenium
from the secondary effluent.
Logsden and Symons (811) researched the removal of mercury
by carbon adsorption and found that powdered carbon, in a jar
test, would adsorb both inorganic and methyl forms in excess of
70 percent.
BIOCIDAL CONTAMINANTS
Carbon adsorption is widely applied to remove organic or
metal-organic biocides. The removal of insecticides and pesti-
cides has been reviewed by Hager and Flentje (523). Dieldrin,
lindane, parathion, and 2, 4, 5-T ester were reduced below the
detectable limit of 0.01 ppb with influent concentrations of
3.6 to 11.4 ppb. Influent concentrations of 3, 5 dinitro-o-
cresol of 30 to 180 ppb were reduced to less than 1 ppb by
carbon adsorption. It was concluded that granular carbo.n beds
will provide a margin of safety for treatment of water contain-
ing varying insecticide or pesticide residues.
Activated carbon removals of several pesticides and PCB's
are well illustrated by results of laboratory studies cited by
the California State Water Resources Control Board (1323), as
shown in Table 39 . A variety of pesticides were experimentally
added to distilled water and passed through carbon filters to
test removal efficiencies. Schwarz (1221) investigated the ad-
sorption of isoprophyl N-(3-chlorophenyl) carbamate (CIPC) onto
activated carbon, concluding that powdered activated carbon
readily adsorbs CIPC from aqueous solution. The adsorption of
CIPC on activated carbon was independent of the pH in the range
of 5 to 9.
Grover and Smith (516) studied adsorption onto activated
carbon of the acid and dimethyl amine forms of 2, 4-D, and
dicambamate. A strong adsorption effect was noted on both the
acidic and salt forms of the compounds. This effect was expec-
ted to increase at low pH values.
On the basis of the literature which has been reviewed, it
can be concluded that activated carbon adsorption is effective
in the removal of some biocidal contaminants; however, further
investigation of this process will be useful.
SYNTHETIC/ORGANIC CONTAMINANTS
As was mentioned earlier, activated carbon and many syn-
thetic compounds are effective at removing organic contaminants
from aqueous solutions, particularly organics of low water solu-
bility. In general, carbon adsorption following secondary treat-
ment is capable of producing an effluent with from 1 to 7 mg/£
of organic carbon (123, 454).
126
-------
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127
-------
Bishop et al . (123) found that carbon adsorption was least
effective in removing highly polar, highly soluble organic
species. DeWalle and Chian (315) found that removal was a
function, at least in part, of the molecular weight of the
synthetic organic contaminant: low (<100) and high (>50,000)
molecular weight compounds are poorly adsorbed. The major
fraction removed by adsorption falls into a 100 to 10,000 mole-
cular weight range.
Studies of the adsorption of 93 petrochemicals by Giusti
et al. (480) confirmed the results of Bishop et al. (123) and
DeWalle and Chian (315): adsorption is largely a function of
molecular weight, polarity, solubility, and branching. Func-
tion was seen to have a substantial effect in conjunction with
solubility and polarity. The relative amenabilities to car-
bon adsorption of straight chain molecules of fewer than four
carbon atoms were as follows: aldehydes >esters >ketones
>alcohols >glycols. For larger molecules, the alcohols moved
ahead of the esters.
Much of the research done on the adsorption of synthetic
organic compounds from wastewater has been concerned with
determining mechanisms of adsorption and optimum removal con-
ditions. The only general removal efficiency studies available
report removals in terms of total organic carbon with little or
no effort made to differentiate the organic compounds involved.
Based on these results, adsorption can reduce the levels of
synthetic organic compounds in a typical domestic wastewater by
75 to 85 percent. If a particular type or types of organic
compounds predominate in a wastewater, these removals must be
adjusted to reflect the effect of compound character on the
adsorption process.
BIOLOGICAL CONTAMINANTS
With the exception of enteroviruses, no information was
found on the adsorption of biological contaminants, although
incidental removal of other organisms would be expected by
filtering action. Adsorption brings about simple removal of
viruses from wastewater rather than inactivation or destruction
(263, 468). Consequently, viable viruses could be reintroduced
to wastewater should desorption of viruses adsorbed to activa-
ted carbon occur.
Columns of granular activated carbon removed between 18
and 40 percent of Type I polio virus from secondary effluent
in studies by Sproul et al. (1313). This research and research
by Gerba et al . (467) using Type I polio virus indicate that
adsorption is inversely related to the concentration of organic
matter in the wastewater. The organics and the virus compete
128
-------
for adsorption sites; consequently, desorption of virus can
occur as adsorption of organic matter continues, or if the
concentration of organics is increased. Several authors have
thus concluded that the process is not dependable for producing
virus-free effluents (246, 468, 1313).
The level of removal actually attained is closely related
to the type of treatment that precedes adsorption. For example,
reducing the concentration of soluble organics in wastewater
by lime coagulation increased polio-virus removal in studies by
Gerba et al. (466, 467). In addition, the degree of virus ad-
sorption from lime-treated wastewater exceeded that from fil-
tered wastewater. These investigations also showed that polio-
virus removal from wastewater effluents is greatly improved by
maintaining a pH value in the range of 3.5 to 4.5. It was
found that virus adsorbed at low pH could become desorbed by a
rise in pH.
129
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TERTIARY TREATMENT: ION EXCHANGE
INTRODUCTION
The process of selective ion exchange has long been util-
ized in the treatment of industrial process waters and in domes-
tic water supply softening. Ion exchange resins (1142) are
classified by the charge of the exchangeable ion. Thus, resins
may be either catonic or anionic. General purpose resins will
selectively exchange both cations and anions. The operational
features of the ion exchange process are well developed and
reliable. Such systems offer a reliable method of removing
inorganic contaminants from the wastewater stream.
As can be seen by an examination of Table 40 , very little
information is currently available on removals of contaminants
from municipal wastewater by use of ion exchange techniques.
The process has not been economically feasible for treatment of
municipal wastewater. Several research programs focusing on
the application of ion exchange to municipal wastewater treat-
ment are presently under way. The most promising future appli-
cation appears to be for ammonia or nitrate removals.
WATER QUALITY PARAMETERS
Eliassen and Tchobanoglous (374, 375) conducted a review
of the literature. They found that removals of phosphorus and
nitrogen by tertiary wastewater treatment incorporating ion
exchange can reach 90 percent. The actual removal efficiency
was seen to depend upon the type of preceding treatment. Evans
(397) investigated the removal of nitrate by ion exchange, con-
cluding that the strong acid/weak base ion exchange process is
well suited for this purpose. With the exception of these few
studies of phosphorus and nitrate, most research performed to
date has focused on ammonium removal, since specific exchange
resins are not available for either the phosphorus or nitrate
ions. However, some zeolite exchange resins do have unusual
selectivity for the ammonium ion. This fact has encouraged
research activity.
On the basis of both pilot and laboratory scale investiga-
tions, it appears that effluent ammonia concentrations of less
than 1 mg/£ can be expected with ion exchange (730, 885). In
a pilot plant study, Mercer et al. (898) used zeolite columns to
test secondary effluent containing 10-19 mg/£ ammonia. Greater
130
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TABLE 40 . LITERATURE REVIEWED PERTAINING TO ION EXCHANGE
Contaminant Reference Number
Water Quality Parameters
Ammonia 4, 120, 165, 196, 277, 374, 454, 730,
885, 898, 1085, 1142, 1501
BOD 119, 120, 196, 277, 487, 885
COD 119, 120, 196, 277, 376, 885, 1547
Chlorides 376, 885
Nitrates 4, 120, 165, 374, 375, 397, 1142, 1501
Nitrites 165, 374, 1501
Phosphates 119, 120, 196, 277, 374, 375, 376, 379,
397, 463, 885
Suspended solids 4, 119, 120, 196
Total dissolved 196, 224, 277, 454, 1547
sol ids
Total organic 119, 120
carbon
Other (general) 531, 622
Elemental Contaminants
Arsenic 520, 1062, 1244
Boron 1140
Cadmium 753, 802, 1062
Chromium 753, 802, 1062, 1157, 1169
Cobalt 753
Iron 753
Lead 753
Manganese 753
Mercury 753, 802, 912, 1062
131
-------
TABLE 40 (continued)
Contaminant Reference Number
Elemental Contaminants
Nickel 753
Selenium 802
Other (general) 760, 1062
Synthetic/Organic 515, 1153
Contaminants
than 99 percent removal of ammonia was achieved. Similarly, 99.7
percent of the ammonia in activated carbon effluent was removed
by a zeolite in laboratory scale experiments by McKendrick et
al . (885).
The Environmental Protection Agency (1085) reviewed pilot
plant studies involving the use of clinoptilolite - a naturally
occurring zeolite - for wastewater treatment. Ammonia removals
ranged from 93 to 97 percent.
It should be noted that the ion exchange process using a
zeolite such as clinoptilolite does not result in the produc-
tion of a sludge containing the removed ammonia. Rather, the
spent zeolite is regenerated with a lime slurry, which is sub-
sequently air stripped, discharging ammonia to the atmosphere.
ELEMENTAL CONTAMINANTS
Ion exchange techniques have been principally applied for
the removal of elemental contaminants from industrial waste
streams (1062). Few studies have dealt with the application of
ion exchange techniques to municipal wastewaters for elemental
contaminant removal. Lindstedtet al. (S02) investigated trace
metal removal and concluded that a cation-anion exchange se-
quence was effective in reducing the concentrations of cadmium,
chromium, and selenium in secondary effluent. Removal effi-
ciencies are summarized in Table 41 .
132
-------
TABLE 41. TRACE METAL REMOVALS BY ION EXCHANGE (802)
Percent Removal After Given Process
Cation Cation-Anion
Trace Metal Exchange Exchange
Cadmium
Chromium
Selenium
99
5
1
99.9
96
99.7
Shen (1244) studied the removal of arsenic from drinking
water by ion exchange, concluding that the process does not
effect adequate removals. Using a variety of input levels, he
was able to achieve arsenic reduction of only 21 percent.
BIOCIDAL CONTAMINANTS
Biocidal contaminant removal through ion exchange has also
received little attention in the literature. In the only study
located, Grover (515) stated that trifluralin, triallate,
diallate, and nonionic herbicides were readily adsorbed on
both cationic and anionic exchange resins, with somewhat more
adsorption occurring on the cationic than on the anionic form.
133
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TERTIARY TREATMENT: NITROGEN REMOVAL PROCESSES
INTRODUCTION
Interim primary drinking water standards established by the
EPA set a nitrate limit of 10 mg/£ in the nitrogen form. Nitro-
gen concentrations in raw municipal wastewaters generally exceed
this value, ranging from 15 to 50 mg/a. Unless facilities are
specifically designed to remove nitrogen, much of it will remain
essentially unaffected, passing through the varying stages of
treatment to ultimately enter the environment. Moreover, reuse
of wastewater treatment plant effluents for direct groundwater
recharge, indirect groundwater recharge through land application,
or indirect reuse as a potable water supply is on the increase.
Such reuse policies make effective nitrogen removal an important
aspect of any wastewater treatment scheme.
In raw municipal wastewater, nitrogen is primarily found in
the form of both soluble and particulate organic nitrogen and as
ammonium ions. Conventional primary and secondary treatment
transforms some of this organic nitrogen into ammonium ions.
Part of the ammonium ion is oxidized to nitrate, and about 15 to
30 percent of the total nitrogen is removed.
Tertiary treatment processes designed to remove wastewater
constituents other than nitrogen often remove some nitrogen
compounds as well. However, removal is often restricted to
particulate forms, and overall efficiency is generally low. Two
tertiary processes particularly designed to remove nitrogen have
been developed: nitrification-denitrification and ammonia
stripping. Tertiary nitrification-denitrification usually
involves two stages. Nitrification occurs in an initial stage,
during which ammonium ions are oxidized to nitrite and nitrate
ions by nitrifying bacteria. These nitrite and nitrate ions are
in turn reduced to nitrogen gas which simply escapes from the
system.
Ammonia stripping is effective only in removing ammonia
nitrogen from municipal wastewater and has no effect on organic
nitrogen, nitric--. ,,r nitrate. Several ammonia stripping plants
are in operation in '.he U.S. (Lake Tahoe, California, Orange
County, California). !>'Jt the process has been found to be
expensive. A nupi>ey r> f te'.nn-ical problems remain to be solved
as well (885),
134
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Nitrification and den i t r i f i cation are biological reactions
which occur naturally during several conventional treatment
processes such as activated sludge treatments, aerobic lagooning,
and anaerobic digestion. The activated sludge process, in
particular, can be closely controlled to promote nitrogen removal.
Such treatment processes are not principally designed to remove
nitrogen, and both nitrification and denitrification occur only
as secondary reactions. For ease of reference, however, all
literature reviewed on the general topic of nitrogen removal has
been tabulated on Table 42 , including tertiary processes
specifically designed for this purpose.
WATER QUALITY PARAMETERS
There is general agreement that a system incorporating
secondary biological treatment and tertiary nitrification-
denitrification should achieve 80 to 95 percent total nitrogen
removal at design flows (531,1001). The nitrification process
alone removes only 5 to 10 percent of the total nitrogen entering
the process, while oxidizing up to 98 percent of the ammonia
nitrogen present to nitrate (458).
Average nitrogen data from systems incorporating nitrifica-
tion-denitrification processes recorded by an EPA Technology
Transfer Publication (458) are presented in Table 43 . Based
on this report, the predicted effluent quality from a nitrogen-
deni trif ication system will be 1.0 nig/2. organic nitrogen, 0.5
mg/£ ammonia nitrogen, 0.5 mg/£ nitrate nitrogen, and 2.0 mg/z
total nitrogen.
Nitrification is in itself an oxygen-demanding process and
therefore reduces the total oxygen demand (TOD) in the waste-
water effluent. Conventional biological or physico-chemical
treatment obtaining up to 90 percent BOD reduction will only
partially reduce the TOD of treated wastewater. For instance,
such treatment will only reduce an influent TOD of 490 mg/a to
an effluent TOD of over 100 mg/£. Nitrification will reduce the
TOD of this effluent to less than 40 mg/£ (458).
Since the denitrification step involves the oxidation of
carbonaceous material, a reduction in biochemical oxygen demand
and total organic carbon can also be expected, in addition to
the effective reduction of TOD.
Nitrogen removal by ammonia stripping was studied by
O'Farrell et al. (1016), who reported a 90 percent removal of
ammonia from a non-nitrified, lime-clarified secondary effluent
at pH 11.5. During a warm weather study performed at Lake
Tahoe, stripping produced a 96 percent removal of ammonia
nitrogen at pH 11.5 and using 400 cu ft of air per gallon of
wastewater (531). Any arbitrary percentage removal can be
achieved with this type of system within available engineering
capabilities, although higher removals mean higher costs.
135
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TABLE 42.
LITERATURE REVIEWED PERTAINING TO
NITROGEN REMOVAL
Contaminant
Water Quality Parameters
Ammoni a
BOD
Chlorides
Nitrates
Phosphates
Suspended
solids
Reference Number
4, 5, 8, 81, 121, 236, 374, 390, 458,
531, 643, 869, 885, 986, 1001, 1016,
1085, 1106, 1142, 1282, 1333, 1351,
1482, 1501, 1546
81 , 458, 646
819
4, 5, 81 , 374, 375, 390, 458, 489,
490, 642, 643, 646, 869, 986, 1001,
1085, 1106, 1142, 1282, 1350, 1351 ,
1501
374
81 , 646, 1350
TABLE 43. EFFLUENT NITROGEN CONCENTRATIONS IN
TREATMENT SYSTEMS INCORPORATING NITRIFICATION - DENITRIFICATION
(458)
Average Effluent Nitrogen, mg/£
Type and Process Sequence Organic-N NH^-N NO^-N NO^-N °N
Lime treatment of raw sewage, 1.1 0.3 0.5 0.0 1.9
nitrification,
denitrification
Primary treatment,
high rate activated
sludge, nitrification,
denitrification,
filtration
Primary treatment,
roughing filters,
nitrification,
denitrification,
filtration
0.8
0.8
0.0
0.9
0.7
0.6
0.0 1.5
0.0 2.3
136
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DISINFECTION: CHLORINATION
INTRODUCTION
Until recently, chlorination was considered virtually an
unmixed blessing as a cheap, effective method to destroy bac-
teria and viruses. It is now recognized, however, that chlori-
nation of wastewater may create chlorinated compounds harmful
to the environment and to human health. The extent of this
potential hazard has not yet been determined; new and existing
wastewater treatment plants continue to utilize chlorine for
disinfection. The primary purpose of municipal wastewater
chlorination is the destruction of pathogenic microorganisms.
This is reflected in the literature reviewed, shown in Table 44.
WATER QUALITY PARAMETERS
Zaloum and Murphy (1553) concluded that chlorination of
treated wastewater effluents does not reduce BOD, COD, and
total organic carbon. Susag (1346), however, found BOD reduc-
tions by chlorination of up to 2 mg/a of chlorine were added.
These values are somewhat misleading, in that BOD reduction
was due both to oxidation of the organic material and to the
formation of chlorinated organics resistant to bacterial action.
When chlorine is added to a wastewater containing ammonia
nitrogen, ammonia reacts with the hypochlorous acid formed to
produce chloramines. Further addition of chlorine converts the
chloramines to nitrogen gas. The reaction is influenced by pH,
temperature, contact time, and initial chlorine-to-ammonia
ratio. If sufficient chlorine is added, 95 to 99 percent of
the ammonia will be converted to nitrogen gas with no signifi-
cant formation of nitrous oxide. The quantity of chlorine
required was found to be 10 parts by weight of chlorine to 1
part of ammonia nitrogen when treating raw sewage. This ratio
decreased to 9:1 for secondary effluents, and 8:1 for lime-
clarified and filtered secondary effluent (1346).
ELEMENTAL CONTAMINANTS
Little information is available on the minimal removal by
chlorination of elemental contaminants. Andelman (27) studied
the effects of chlorination on barium, copper, and nickel. The
treatment effected a 34 percent reduction in barium, a 5 percent
reduction in nickel, and had no effect upon copper. Kokoropoulos
137
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TABLE 44 . LITERATURE REVIEWED PERTAINING TO CHLORINATION
Contaminant Reference Number
Water Quality Parameters
Ammonia 43, 76, 114, 120,127,727, 728, 1085,
1116, 1288, 1555
BOD 120, 321, 593, 1046, 1084, 1215, 1346,
1553
COD 120, 1084, 1091, 1346, 1553
Chlorides 684
Cyanides 65, 1046
Nitrates 120, 684, 1085, 1322
Nitrites 727, 728, 1085
Phosphates 120, 684
Suspended solids 120, 1084, 1215
Total dissolved 1215
solids
Total organic 120, 1553
carbon
Others (general) 385, 391, 531, 727, 728, 884, 1468
Elemental Contaminants
Barium 27
Boron 1367
Copper 27
Iron 727, 1046
Manganese 727, 728, 1181
Mercury 1046, 1499
Nickel 27, 728
138
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TABLE 44 (continued)
Contaminant Reference Number
Synthetic/Organic 5, 84, 158, 191, 482, 657, 971, 1130,
Contaminants 1250, 1287, 1332, 1450, 1555
Biological Contaminants
Adeno virus 95
Bacteria 8, 79, 161, 210, 213, 283, 292, 321, 395,
615, 651, 830, 1304
Coliforms 161, 164, 292, 321, 395, 383, 395, 483,
612, 615, 727, 728, 937, 1084, 1304,
1312, 1389, 1393, 1458, 1489, 1490
Coxsackie virus 95, 257, 380, 615, 890, 1206
(A & B)
ECHO virus 380, 615, 1260
Escherichia coli 141, 375, 383, 428, 615, 1206, 1414, 1430
Fecal streptococci 292, 395, 612, 1084
Hepatitis virus 615
Mycobacterium 428, 615
Parasitic worms 161, 615, 1188, 1304
Polio virus 161, 223, 380, 615, 817, 820, 1206, 1260,
1293, 1425
Protozoa 161, 383, 428, 1335
Salmonella 161, 292, 395, 428, 1083
Shigella 161
Vibrio cholerae 161
Virus 79, 92, 95, 96, 100, 101, 127, 161, 210,
213, 280, 281, 283, 352, 382, 383, 395,
454, 492, 615, 651, 739, 745, 816, 817,
900, 970, 1009, 1083, 1095, 1243, 1259,
1260, 1304, 1366, 1434
Other (general) 161, 588
139
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(727) reported hypochlorous acid reacted with soluble iron (II)
and manganese (II) to form precipitates.
SYNTHETIC/ORGANIC CONTAMINANTS
No research was found to address removal or destruction
effects of chlorination on any of the synthetic/organic contami-
nants. However, considerable interest has recently developed
concerning the production of chlorine-containing organic com-
pounds by chlorination. The reactions of chlorine with organic
compounds in water are diverse, including oxidation, substitu-
tion, addition, and free radical reactions. Chlorination may
produce several different chlorinated products from a single
organic pollutant molecule. Some of these compounds have been
identified as toxic to aquatic life by Snoeyink (1287), Brungs
(158) , and others.
Jolley (657) evaluated chlorine-containing organic consti-
tuents in chlorinated effluents and found that stable chlorine-
containing compounds were present after effluents had been
chlorinated to a 1 to 2 mg/£ chlorine residual. These compounds
are identified in Table 45.
TABLE 45. IDENTIFICATION OF CHLORINE CONTAINING
CONSTITUENTS IN CHLORINATED EFFLUENTS (657)
5 - Chlorouracil 5 - Chlorouridlne
8 - Chlorocaffeine 6 - Chioroguanine
8 - Chloroxanthine 2 - Chlorobenzoic acid
5 - Chlorosalicylic acid 4 - Chloromandelic acid
2 - Chlorophenol 4 - Chlorophenylacetic
acid
4 - Chlorobenzoic acid 4 - Chlorophenol
3 - Chlorobenzoic acid 3 - Chlororesorcinol
3 - Chloro-4-hydroxybenzoic acid
4 - Chloro-3-methylphenol
A similar project was conducted by Glaze and Henderson
(482). The chlorinated organics identified in this study are
listed in Table 46.
Shimizuetal. (1250) stated that halogenated nucleic acid bases
are incorporated into the nucleic acid. Also, the incorporation
of 5-deoxybromouridine in DNA and 5-fluorouraci1 into RNA are
known to cause mutations. No work has been completed to deter-
mine how nucleic acids react with chlorine or the resulting
mutations.
140
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TABLE 46. CHLORINATED OR6ANICS IN
WASTEWATER EFFLUENT (482)
Chloroform
Dichlorobutane 3-chl
Chlorocyclohexane (-18)
0-di chlorobenzene
P-dichlorobenzene
Pentachloroacetone
Trichlorobenzine
Chiorocumene
N-methyl-tri chloroani1i ne
Trichlorophenol
Chloro-a-methyl benzyl alcohol
Di chloromethoxytoluene
Trichloromethylstyrene
Dichloro-a-methyl benzyl alcohol
Dichloro-bis (ethoxy) benzene
Dichloro-a-methyl benzyl alcohol
Trichloro-a-methyl benzyl alcohol
Trichloro-a-methyl benzyl alcohol
Tetrachloroethylstyrene
Tetrachloromethoxytoluene
Dichloroani1ine derivative
Dichloroani1ine derivative
Trichlorophthalate derivative
Tetrachlorophthalate derivative
Dibromochloromethane
3-chloro-2-methylbut-l-ene
Chloroalkyl acetate
Tetrachloroacetone
Chloroethylbenzene
Hexachloroacetone
Dichloroethyl benzene
Dichlorotoluene
Trichloroethyl benzene
Tri ch1oro-N-methylan 1 sole
Tetrachlorophenol
Trichlorocumeme
Trichlorodimethoxybenzene
Dichloroacetate derivative
141
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BIOLOGICAL CONTAMINANTS
The effectiveness of chlorination as a disinfection process
has long been recognized. All researchers are in agreement that
the effectiveness of disinfection by chlorine is influenced by
time and chlorine concentration and also by: (1) whether the
chlorine residual is free or combined; (2) effectiveness of
mixing; (3) whether or not particulates are present; (4) pH;
(5) temperature; and (6) the concentration, condition, and nature
of the organisms. Keeping these limitations in mind, an idea
of the relative resistances of organisms to disinfection by
chlorine can be seen in Table 47.
TABLE 47. EFFECT OF CHLORINATION ON VARIOUS ORGANISMS (615)
Group
Virus
Bacteria
Nematodes
Organism
Infectious
hepatitis
Coxsackie
Coxsackie
Echo
Poliovirus I
Coliphage B
Theiler Phage
M. tuberculosis
E. coli
Col iforms
Total Count
Di pi ogaster
Cheilobus
Chlorine
Residual
(mg/£)
1
15
5
1.0
1.95
0.53
0.03
0.03
1-5
2
1
0.14
0.03
1-1.2
some
2.5-3
15-45
Time
min.
30
30
2.5
3
6.5
14
10
10
120
30
30
3
10
15
15
120
1
Efficiency
Survived
Inactivated
Survived
99.6% Inactivated
Survived
Survived
20% Survival
Inactivated
99% Kill
99% Kill
Destroyed
99.9% Kill
52% Kill
99% Kill
98-99% Kill
Survived and
Mobi le
Others
S. mansom
Vo v a and
mi racidia)
S. japonicum
(ova and
m i r a c i d i a)
0.2-0.6
0.2-0.6
30
30
Killed
Killed
Eliassen and Tchobanoglous (375) found that 2 to 6 mg/£ of
chlorine applied for 20 min would effect a 99.99 percent kill
of the total coliforms, fecal coliforms, and fecal streptococci
present in wastewater influent.
142
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The effect of chlorination on entamoebic cysts was investi-
gated by Stringer and Kruse (1335). It was concluded that the
hypochlorous acid (HOC1) form of free available chlorine was
the most rapidly acting cysticide compared with other halogen
species. Entamoeba histolytica cysts and tapeworm eggs with-
stand the ch1 ori nation treatment generally applied to waste
treatment effluent (161). Rudolfs et al. (1188) reported the
development of active embryos from a majority of ascaris eggs
in contact with 5,000 ppm chlorine solutions for 30 min.
Davis and Keen (292) conducted a research project on the
ability of municipal wastewater bacteria to survive chlorination
and reestablish populations after discharge. Fecal coliform,
fecal streptococci, and total coliform enumeration was performed
with further differentiation into lactose nonfermenters within
and outside the family Enterobacteriaceae. It was determined
that the vast majority of the wastewater bacterial species that
reestablished their populations within 21 days following chlori-
nation were lactose nonfermenters not included in the Entero-
bacteriaceae. Many of these bacteria could be pathogenic under
the appropriate conditions and may constitute a threat to public
health in receiving waters designated for contact recreation.
Virus inactivation is one of the more difficult tasks of
chlorination, but as in any disinfection process, required kills
can be achieved by lengthening the time or increasing the con-
centration (615). A study of the inactivation of viruses in
wastewaters by chlorination was performed by Lothrup and Sproul
(817). It was ascertained that:
t High-level inactivation of viruses can be obtained in
treated and untreated domestic wastewaters. Present
chlorination practices (1 mg/£ of residual), however,
are inadequate for a high level of virus inactivation.
• A combined chlorine residual of 28 mg/£ was required
to produce a 99.99 percent inactivation of the T2
bacteriophage in settled raw wastewater after a 30-min
contact time.
• A combined chlorine residual of 40 mg/£ was required to
provide a 99.99 percent destruction of the Type I
polio virus in settled wastewater after a 30-min contact
time.
t Free chlorine residuals of 0.2 to 0.4 mg/£ , after 30
min, produced a complete inactivation of the polio virus
and T2 phage in the secondary effluent.
143
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• In experimental runs with synthesized storm-water over-
flow samples, a 100 percent inactivation of the Type I
polio virus was obtained by providing a free chlorine
resi dual .
• The T2 bacteriophage was much less sensitive to combined
chlorine residuals than is the coliform organism and
somewhat more sensitive than the polio virus to combined
chlori ne residuals .
A similar research project investigating the inactivation
of enteroviruses by chlorination was conducted by Shuval et al .
(1260). The following conclusions were obtained:
t The strain of ECHO virus used was sensitive to the com-
bined chlorine in the sewage, with reductions of 99 per-
cent in 30 min and 99.93 percent in 6 hr using 3.6 mg/£
of chlorine. No virus was recovered in the sample tested
after 4 hr contact with 7 mg/£ of chlorine dosage or
after 2.5 hr with 11 mg/l of applied chlorine. Inacti-
vation was shown to be a function of time and chlorine
concentration under the test conditions.
• The strain of polio virus Type I used was much less sen-
sitive to the combined chlorine, showing only 50 and 90
percent reductions in 6 hr using 4 and 11 mg/£ chlorine
dosage respectively.
• Coliform reductions obtained followed known patterns,
with a standard of less than 100 coliforms/100 mi being
obtained in 80 percent of the samples after 2 hr of
contact with a chlorine dose of about 8 mg/£.
• Although the inactivation of the ECHO virus strain was
comparable to that obtained with coliform organisms,
this study indicates that the coliform index in chlori-
nated sewage may not give a true picture of the degree
of inactivation obtained with the more resistant strains
of enterovirus such as polio virus Type 1.
The minimum concentration of chlorine required for complete
inactivation of the Sabin oral poliovaccine Type I virus strain
was examined by Varma et al. (1425). Various exposure periods
with pH 5.2 at 20°F were studied: a concentration of 22 mg/£ for
5 min of exposure time, 19 mg/£ for 15 min, 19 mg/l for 30 min,
17 mg/£ for 45 min, and 14 mg/£ for 60 min. Nonetheless, on
the basis of the literature surveyed, it is evident that chlori-
nation per se does not provide conclusive proof of disinfection.
Boardman and Sproul (127) described the protection afforded
viruses associated in particulate matter. Surface adsorption
144
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gave no viral protection. When viruses are embedded within
particles, the disinfecting molecules must diffuse through the
particle matrix before reaching the virus and initiating any
inactivation. Chemical diffusion is a slow process; as a con-
sequence, virtually all the embedded viruses are prtected from
disinfection.
145
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DISINFECTION: OZONATION
INTRODUCTION
Ozone, an allotropic form of oxygen, is a powerful oxidizing
agent for the disinfection of wastewater. Ozone is used in over
100 municipalities in Europe for disinfection of drinking water.
Certain chemical features make ozone treatment a particularly
attractive method of water purification:
t It is a powerful oxidant which reacts rapidly with most
organic compounds and microorganisms in wastewater.
• It does not impart taste and odor to potable water.
• It is produced from oxygen in air by means of electric
energy.
On the negative side, the cost of ozonation is not presently
competitive with chlorine disinfection. Moreover, long-term
residual disinfection capabilities are lacking, and the insta-
bility of ozone generally necessitates its generation on site
(1429).
The principal ozone decomposition products in aqueous
solution are molecular oxygen and the highly reactive free
radicals OH, H02, H0,+. Very little is known about the sig-
nificance of th'e free radical intermediates on the germicide!
properties of ozone solutions. The same free radicals are pro-
duced by irradiation of water, and it has been reported that
H02 and OH. radicals contribute significantly to the killing of
bacteria by this process.
As seen in Table 48, a considerable amount of information is
available on the destruction of various pathogens by ozonation,
however little information was found on the effect of ozone
upon other contaminants.
WATER QUALITY PARAMETERS
Because of its strong oxidizing character, ozone is very
reactive toward the organic compounds which make up the BOD,
COD, and total organic carbon. Under ideal conditions the
reactions would result in almost complete oxidation and only
carbon dioxide as a reaction product. In practice, ozonation results
146
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TABLE 48. LITERATURE REVIEWED PERTAINING TO OZONATION
Contami nant
Water Quality Parameters
Ammonia
BOD
COD
Nitrates
Nitrites
Phosphates
Suspended solids
Total organic
carbon
Others (general )
Elemental Contaminants
Adeno virus
Bacteria
Clos tridi urn
botulinium
ECHO virus
Escherichia coli
Fecal streptococci
Parasitic worms
Polio virus
Salmonel1 a
Virus
Other (general)
Reference Number
8, 470, 651, 709, 1166
320, 321, 709, 868, 970, 985
236, 320, 470, 707, 985, 1231
470, 651, 985, 1166
470, 651, 709, 1166
985
470
985, 1262
394
437, 438
95, 213
8, 72, 213, 320, 321, 469, 542, 707,
709, 1178
1429
1429
72, 95, 686, 687, 707, 1429
72, 707
897, 1429
213, 223, 686, 687, 835
707
95, 96, 97, 213, 280, 320, 492, 816,
818, 1069, 1070
686
147
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in only partial oxidation and produces simpler organic molecules.
Both Ghan and Nebel reported COD removals of Less than 40
percent. Morris found that the apparent BOD of a wastewater can
increase after ozonation as a result of refractory organic
molecules being oxidized to simpler, biodegradable compounds.
If the ozone is applied after other treatment processes (as is
normal), the increase in organic nutrient molecules can lead
to the growth in the distribution system of algae, slime
bacteria, and the possible regrowth of any pathogens not des-
troyed during treatment.
Ozone is effective at decreasing concentrations of organic
suspended solids and organic nitrogen through oxidation.
Ozonation can assist in suspended solids removal through froth
flotation mechanisms induced through the process. Ozone will
also oxidize ni-trites to nitrates, but will not react with
ammonia (651). There is little evidence to date that ozonation
will produce any toxic or carcinogenic oxidation by-products as
will chlorination.
ELEMENTAL CONTAMINANTS
Furgason and Day (438) studied the feasibility of
ozonation for iron and manganese removal from raw water with
relatively high input concentrations. The study demonstrated
that ozone effectively oxidized the iron and manganese to
an insoluble form which could be filtered from the water.
Complete oxidation of the minerals required a reaction time
of 30 sec. Filtration studies indicated a relatively fine
medium was required to remove the oxide precipitate.
BIOLOGICAL CONTAMINANTS
The use of ozone as a wastewater disinfectant was reviewed
by Venosa (1429). It was concluded that with 0.1 mg/2- of active
chlorine, 4 hr would be required to kill 6 x 104 E. coli cells
in water, whereas with 0.1 mg/a of ozone only 5 sec would be
necessary. When the temperature was raised from 22°C to 37°C,
the ozone inactivation time decreased from 5 sec to 0.5 sec.
These investigations revealed that the contact time with ozone
necessary for 99 percent destruction of E. coli was only one-
seventh that observed with the same concentration of hypochlo-
rous acid. The death rate for spores of Bacillus species was
about 300 times greater with ozone than with chlorine.
In the same study, Venosa also described bacteriological
studies performed on secondary effluent from an extended aera-
tion pilot plant in the Metropolitan Sewer District of Louisville,
Kentucky. Using an average applied ozone dosage of 15.2 mg/£
for an average contact time of 22 min, fecal coliform reductions
of greater than 99 percent were achieved, resulting in a mean
148
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fecal coliform concentration of 103 cells/100 ma , a mean
total coliform concentration of 500 cells/100 ma , and a mean
fecal streptococci concentration of 8 cells/100 m£ in the
final effluent. Laboratory results with raw sewage indicated
that ozone could be successfully used to sterilize sewage
containing Bacillus anthracis, influenza virus, and B. subti1 is
morph. globibii, and to inactivate toxins of Clostri dium
botuli num. Ozone consumption was 100 to 200 mg/£ for 30 min.
Finally, Venosa found ozone to be many times more effective
than chlorine in inactivating poliomyelitis virus. Identical
dilutions of the same strain and pool of virus, when exposed
to 0.5 to 1.0 mg/a of chlorine and 0.05 to 0.45 mg/Ji of ozone,
were devitalized within 1.5 to 2 hr by chlorine, while only
2 min of exposure were required with ozone.
Majumdar et al. (835) also studied the inactivation of
polio virus by ozonation and concluded that the inactivation
is not complete. Results are summarized in Table 49.
TABLE 49. SURVIVAL OF POLIO VIRUS IN
OZONATION CONTINUOUS FLOW STUDIES (835)
Ozone Residence Average
Type of Concentration Time Survival
Wastewater (mg/£) (min) (percent)
Primary
wastewater 0.84 8.0 1.820
1.47 2.0 0.016
4.44 1.0 0.006
Secondary
wastewater 0.79 8.0 2.055
1.77 2.0 0.013
5.05 1.0 0.006
Pavoni and Tittlebaum (1069) recently studied ozone
disinfection of viruses in the Fort Southworth Pilot Plant
of the Metropolitan Sewer District in Louisville. Using
F2 bacteriophage as the model virus, they demonstrated virtually
100 percent inactivation efficiency in the secondary effluent
after a contact time of 5 min at a ozone dosage of approximately
15 mg/x, and a residual of 0.015 mg/£ . Of particular interest
was the observation that the rate of inactivation was greater
for F2 bacteriophages than bacteria. In addition, the following
conclusions were reached:
1. F2 virus concentrations were shown to be unaffected
by the flow or mixing of the ozone reactor.
149
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2. F2 virus was inactivated with virtually 100 percent
efficiency after a contact time of 5. min at a total
ozone dosage of approximately 15 mg/£ and a residual
of 0.015 mg/a .
3- E. coli bacteria and F2 virus were inactivated with
virtually 100 percent efficiency after a contact time
of less than 15 sec in the absence of ozone-demanding
material.
4. An extremely small number of viral particles was
observed in effluent studies.
5. Oxidation by ozone appears to be the mechanism of
kill for bacterial cells and viral particles. Ozone
is theorized to act as a general oxidant causing
cell lysis and the release of soluble COD.
Mercado-Burgos et al. (897) examined the effect of ozone
on Schistosoma ova, concluding that the process was ineffective
150
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SLUDGE TREATMENT AND DISPOSAL SYSTEMS
THICKENING AND DEWATERING
INTRODUCTION
Sludge is the most notable by-product of the wastewater
treatment plant. Disposal of sludge materials is a problem
comparable in magnitude and importance to that of wastewater
treatment. Because of the high water content of sludge, any
reduction in volume through dewatering or thickening, of these
by-products marks a significant step towards an improved solu-
tion of the disposal problem. Processes available for sludge
volume reduction include:
• Gravity sludge thickeners
• Air flotation sludge thickeners
• Centrifugal sludge thickeners
• Vacuum fi1ters
• Drying lagoons
• Pressure filtration
Sludge dewatering by land methods (e.g., lagooning) has
been the accepted practice for many years. However, attention
has recently focused on the use of other techniques available
for dewaterina. Increased land costs together, with stringent
land, air, and water pollution standards have made sludge
dewatering an economically attractive process since it sig-
nificantly reduces the volume of sludge requiring disposal.
Literature concerning sludge thickening and dewatering
processes, as reviewed in Table 50 , is rarely pertinent to
discussions of adverse health effects of sludge-contained
contaminants. In general, this literature deals mainly with
engineering design and operation parameters and the relative
efficiency of dewatering processes, with only slight coverage
of partitioning, degradation, or concentration of either
general, elemental, or biological contaminants.
Coagulants are used for improved dewatering characteris-
tics and have been indicated as possible factors providing
viral inactivation or removal (246). Kampelmacher and Jansen
(670) reported reduction of bacteria in sludge through vacuum
filtration and chemical conditioning by ferric chloride,
151
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TABLE 50 . LITERATURE REVIEWED PERTAINING TO
SLUDGE THICKENING AND DEWATERING
Contamlnant Reference Number
Water Quality Parameters
Ammonia 925, 1080
BOD 16, 201, 334, 1435
COD 16, 334
Chlorides 1080
Nitrates 334
Phosphates 546, 1080
Suspended 16, 201, 229, 438, 518, 546, 864, 1008,
solids 1015, 1054, 1058, 1080, 1123, 1308, 1368,
1369, 1435
Others (general) 16, 201, 334, 401, 518, 531, 546, 551,
552, 622, 638, 639, 764, 1008, 1015,
105J, 1054, 1055, 1058, 1123, 1308, 1369,
1397
Elemental Contaminants
Aluminum 246, 1080
Boron 1080
Cadmium . 229, 1080
Chromium 229, 1080
Copper 229, 635, 1080
Iron 229, 1080
Lead 1080
Manganese 229, 1080
Mercury 635
Nickel 635, 1080
152
-------
TABLE 50 (continued)
Contami nant Reference Number
Zinc 229, 635, 1080
Other (general) 229
Biological Contaminants
Bacteria 670
Coliforms 401 , 568
Coxsackie virus 1053
(A & B)
Escherichia coli 568
Salmonella 244, 670
Virus 246
ferrous sulfate, and lime. Two separate analyses of Dutch
treatment plants showed a marked reduction (from two to
four orders of magnitude) of aerobic bacteria and entero-
bacteriaceae.
Peterson et al. (1080) studied the chemical properties of
waste activated sludge after vacuum filtration; his results,
presented in Table 51, provide a representative picture of
the constituents found in thickened and dewatered sludge.
153
-------
TABLE 51 . CHEMICAL PROPERTIES OF
VACUUM FILTERED WASTE ACTIVATED SLUDGE* (1080)
Analyses % of Dry Wt Basis
Total Nitrogen 6.37
Ammonia Nitrogen Trace
Phosphorus 2.49
Zinc
Boron 0.002-0.04
Iron 5.32
Manganese 0.012
Aluminum
Cadmium 0.028
Chloride
Chromium 0.362
Copper 0.11
Nickel 0.034
Lead 0.141
* Waste activated sludge; ferric chloride addition;
vacuum filtered; heat dried
154
-------
ANAEROBIC DIGESTION
INTRODUCTION
Sludge digestion converts bulky, odorous, raw wastewater
sludge to a relatively stable material that can be readily de-
watered Tfid disposed without excess obnoxious odors. Bacterial
digestion is a two-stage process effecting the decomposition
of organic and/or inorganic material in the absence of free
oxygen. The first stage is accomplished by facultative
bacteria, which break down large organic compounds and con-
vert them to simpler organic acids. Acid-splitting, methane-
forming, anaerobic bacteria complete the second step, con-
verting the organic acids to methane and carbon dioxide.
These strict anaerobes have a relatively slow growth rate
and are highly sensitive to environmental conditions of
temperature, pH, and anaerobiosis.
Health-related problems traceable to anaerobic digestion
would be dependent upon the final sludge solids disposal scheme,
since the dewatering liquor circulates inaclosed loop system
within the treatment plant. Pertinent literature reviewed con-
cerning this subject is indicated in Table 52.
WATER QUALITY PARAMETERS
Sekikawa et al. (1230) investigated the solubi1ization of
certain water quality parameters under anaerobic conditions.
A period of marked increase in dissolved phosphate concentra-
tion and BOD value was noted during the initial stage of
digestion. In the ensuing stage, these values decrease. The
maximum amount of phosphorus released equaled approximately
50 percent of the total phosphorus contained in the sludge.
These temporary increases were attributed to autolysis, or
microbial destruction, of sludge organisms. Table 53 shows
typical anaerobic digester content characteristics during the
study.
ELEMENTAL CONTAMINANTS
Anaerobically digested sludge characteristically contains high
concentrations of elemental contaminants. Sludge volume is
reduced during digestion, with a corresponding increase in
metal concentrations. The extent to which these metals inhibit
the digestion process, however, has in the past received more
attention than any ultimate effect on human health.
155
-------
TABLE 52 . LITERATURE REVIEWED PERTAINING
TO ANAEROBIC DIGESTION
Contaminant Reference Number
Water Quality Parameters
Ammonia 580, 584, 690, 864, 1080, 1190
BOD 205, 1080, 1230
COD 1080
Chlorides 1080
Nitrates 580, 584, 1195
Nitrites 112, 584
Phosphates 513, 580, 584, 623, 690, 736, 838, 1035,
1036, 1080, 1195, 1228, 1230, 1267
Suspended 497, 838, 1035, 1054, 1055, 1080, 1128,
solids 1230
Others (general) 531, 551, 552, 622, 690, 1035, 1053,
1123, 1230
Elemental Contaminants
Aluminum 513, 690, 1035, 1036, 1080, 1228
Barium 690
Boron 690, 1180, 1195
Cadmium 228, 580, 584, 690, 853, 975, 1063, 1080,
1195, 1400
Chromium 71, 228, 580, 584, 623, 690, 853, 1080,
1195, 1400
Cobalt 1195, 1400
Copper 71, 228, 584, 623, 690, 761, 853, 975,
1063, 1080, 1195, 1400, 1411
Iron 112, 228, 580, 690, 761, 975, 1080, 1228,
1267
156
-------
TABLE 52 (continued)
Contami nant Reference Number
Lead 228, 580, 584, 690, 853, 1063, 1080,
1195, 1400
Manganese 112, 228, 580, 584, 623, 690, 1080, 1195,
1400
Mercury 228, 690, 803, 1195
Nickel 71, 228, 584, 623, 690, 761, 1063, 1080,
1195, 1400
Zinc 71, 112, 228, 401, 584, 623, 690, 761,
853, 975, 1063, 1080, 1195, 1228, 1400
Other (general ) 1228
Biocidal Contaminants
DDT 14, 15
Synthetic/Organic
Contaminants 714
Biological Contaminants
Bacteria 161, 580, 935, 1123
Coliforms 568, 580, 582, 935
Coxsackie virus 95, 112, 161, 428, 1053
(A & B)
ECHO virus 112
Escherichia coli 568, 1080, 1123
Mycobacteriurn 428
Parasitic worms 161, 428, 791
Polio virus 112, 161, 428
Protozoa 428, 1080
Salmonella 95, 161, 428, 717, 1080, 1123
Shigella 161
157
-------
TABLE 52 (continued)
Contaminant Reference Number
Vibrio cholerae 161
Virus 95, 96, 112, 161, 428, 580, 908, 1209,
1476
Other (general ) 161
TABLE 53 . SLUDGE CHARACTERISTICS (1230)
Parameter Concentration
Sludge volume index (%) 25.0
Total suspended solids (mg/£) 2,784
Volatile suspended solids (mg/£) 1,920
Ash (mg/£) 864
Total nitrogen (mg/£) 282
Total phosphorus (mg/£) 60
as Phosphate 183.5
BOD (mg/£) 1,465
Table 54 presents the concentration range of various
constituents including elemental contaminants found in
anaerobically digested liquid sludge from 35 Wisconsin munici-
palities, recorded by Keeney et al. (690). Table 55 shows
the average metal concentrations reported by Salotto et al.
(1195). The authors thought the geometric mean would be the
best measure of central tendency. The spread associated with
the geometric mean was greater for certain metals such as
cadmium, chromium, and manganese, indicating that these
metals -- not ordinarily found in domestic wastewater -- may
occasionally be introduced (probably from industrial sources)
in relatively high concentrations.
A laboratory investigation of mercury distribution in an
anaerobic digester was reported by Lingle and Hermann (803).
Data indicate that during experimentation, more than 96 percent
of the mercury from phenyl mercuric chloride and mercuric
chloride remained in the sludge solids. It is suggested that
the biological methylation of mercury may have been inhibited
by the presence of sulfides in the digester sludge. The
158
-------
TABLE 54 . CONSTITUENT CONCENTRATIONS IN
ANAEROBICALLY DIGESTED SLUDGE (690)
Concentration
Constituent Range*
Total nitrogen (moist) 3.4-9.5
Total nitrogen (dried) 2.4-3.1
Ammonia nitrogen (moist) 0.8-4.1
Ammonia nitrogen (dried) 0.02-0.26
Organic Carbon 25.7-38.5
Phosphorus 2.7-6.1
Aluminum 0.36-1.2
Iron 0.8-7.8
Cadmium/Zinc 0.15-33
Zinc 490-12,220
Copper 140-10,000
Nickel 15-1,700
Cadmium 5-400
Lead 40-4,600
Chromium 50-32,000
Silver 0.6-31
Boron 150-750
Manganese 180-1 ,130
Barium 530-1 ,340
Strontium 52-7,810
* Range for the first 9 constituents is given in percent
of solids and in mg/kg for the last 11 constituents.
159
-------
TABLE 55. AVERAGE CONCENTRATIONS OF METALS IN DIGESTED SLUDGE (1195)*
Arithmetic
Metal
Silver
Boron
Cadmium
Calcium
Chromium
Cobalt
Copper
Mercury
Manganese
Nickel
Lead
Strontium
Zinc
Mean
250
430
75
36,500
1,860
350
1,590
10
1,300
680
2,750
520
4,210
Std. Dev.
(+ and -)
230
310
104
23,800
1,920
220
1,670
18
2,290
620
2,350
670
3,800
Geometries
Mean
190
380
43
31,100
1,050
290
1,270
6.5
475
530
2,210
290
2,900
Std. Dev.
(f and x)
1.99
1.58
2.47
1.77
3.22
1.88
1.95
2.34
3.67
1.88
1.82
2.70
2.40
Median
5%
Value
100
350
31
30,000
1,100
<100
1,230
6.6
380
410
830
175
1,780
* (All figures mg/kg dry sludge basis)
160
-------
sulfide content of the sludge (228 mg/£) may have precipitated
the mercury, thus making it unavailable for biological methyla-
t i o n .
The use of sulfide as a detoxicant for other heavy metals
has also been suggested, with the metal ions being precipitated
out as metallic sulfides (552).
BIOLOGICAL CONTAMINANTS
The anaerobic digestion process effectively degrades
proteins and other organic materials. Many studies have been
made of the destruction or inactivation of viruses and bacteria
Hinesly et al. (580) reported the incidence of fecal coliforms
to be approximately 10^ cells/m£ of digested sludge. The
number of fecal coliforms in liquid digested sludge gradually
decreases with increased digester retention time, as shown in
Table 56.
SI
udge
Total
SI
udge
LI
TABLE
QUID D
56. NUMBER OF
IGESTED SLUDGE
sampl es
si
s
udge
uperna
tant
0
4x1
3x1
FECAL COLI
AS A FUNCTI
0
0
4
3
Day
19
7x1
2x1
FORMS
ON OF
s
O3
O1
PER m*
TIME (
32
2xl02
0
OF
580)
A study of the destruction of various species of bacteria
by anaerobic digestion was performed for the EPA (1123).
Results of the study (presented in Table 57 ) indicate that
certain organisms were not entirely eliminated by digestion.
This was corroborated in a review by Foster and Engelbrecht
(428). The authors reported that destruction of S_. typhpsa
was 83 to 99.7 percent after 12 hr in batch digestion studies,
A 31 percent survival of mycobacteriurn species was noted after
digestion of more than 35 days. Tubercle bacilli were able
to survive from 6.5 months up to 2 years after digestion,
even though 90 percent of the organisms were destroyed during
treatment,
Viral inactivation by anaerobic digestion is accomplished
far more rapidly than bacteria destruction. The mechanism
of infectivity loss or inactivation has not been determined,
although the linearity of inactivation curves suggests that
a single mechanism predominates (112).
161
-------
Anaerobically digested sludge seeded with a swine entero-
virus (EPCO-1) was fed to germ-free piglets in experiments
performed by Meyer et al. (908). Indications of piglet
infection were found after four but not five days of digestion,
Based on the data provided, an inactivation rate of 91 percent
per day can be estimated.
The inactivation of polio virus I, coxsackie virus A-9,
coxsackie virus B-4, and ECHO virus II was studied by
Bertucci et al . (112). According to this study, viral inac-
tivation appeared to follow a first order reaction pattern.
Significant differences were noted among the inactivation
rates of the virus tested.
TABLE 57. BACTERIAL DESTRUCTION BY ANAEROBIC DIGESTION
(1123)
Bacteria
Endamoeba
hystolytica
Salmonella
typhosa
Tubercle
bacilli
Digestion Period
(days)
12
20
35
Destruction
%
<100
92
85
Remarks
Greatly reduced pop-
ulations at 68°F
85% reduction in 6
days detention
Digestion cannot be
relied upon for com-
plete destruction
Escherichia
coli
49
<100
Greatly reduced pop-
ulations at 99°F,
about the same re-
duction in 14 days
at 72°F
162
-------
AEROBIC DIGESTION
INTRODUCTION
Aerobic digestion is another method used to stabilize the
organic fraction of wastewater treatment plant sludge. Avail-
able literature is listed on Table 58. To date, this pro-
cess has only been employed at small wastewater treatment
plants or where polishing of excess waste-activated sludge is
necessary. Widespread application of aerobic digestion is
limited by the long solids retention time and consequential
high oxygenation costs required by the process.
Aerobic digestion is a biological process during which
oxidation is completed in two phases: (1) direct oxidation
of any biodegradable matter by microorganisms; and (2) oxi-
dation of microbial cellular material by endogenous respira-
tion. After input waste organics are completely oxidized,
the sludge mass is further reduced by endogenous metabolic
activity.
Advantages of aerobic digestion include: (1) a volatile
solids reduction comparable to that obtained by anaerobic
digestion; (2) a humus-like digested sludge that has no dis-
agreeable odor; (3) a relatively low BOD concentration in the
supernatant; (4) a sludge that is easily dewatered; and
(5) a relatively problem-free operation (531, 552, 622).
However, the process has certain disadvantages: power costs
to supply the required oxygen are high, and pH values as low
as 4.5 may result due to the absence of methane gas production.
Such extreme pH values inhibit nitrate formation and result
in the release of large amounts of soluble phosphorus (552).
163
-------
TABLE 58. LITERATURE REVIEWED
PERTAINING TO AEROBIC DIGESTION
Contaminant Reference Number
Water Quality Parameters
Ammonia 327
BOD 531, 552, 622, 1053
COD 456
Nitrates 327, 1281
Phosphates 552, 1230
Suspended 456, 531, 552, 622, 1053
solids
Other (qeneral) 531, 551, 552, 691, 1053, 1123
Elemental Contaminants
Chromium 71
Copper 71
Nickel 71
Zinc 71
Biological Contaminants
Bacteria 1281
Coliforms 568
Escherichia 568
coli
Virus 1281
164
-------
THERMAL PROCESSES
INTRODUCTION
Thermal processes, whether conventional heat treatment or
incineration, use increased temperature to achieve either
pasteurization or oxidation of organic material. Literature
concerning these processes is shown in Table 59. Reports on
heat treatment mainly cover the Inactivation of biological
contaminants; those on incineration more frequently address the
subject of elemental contaminants.
Heat treatment is a well known method of destroying patho-
genic organisms (1123). This wet air oxidation process is based
on the principle that any substance capable of burning can be
oxidized in the presence of liquid water at temperatures between
250°F and 700°F. The process can operate on difficult-to-dewater
waste liquors and sludges with extremely low solids concentra-
tion. In general, given the proper temperature, pressure,
reaction time, and sufficient compressed air or oxygen, any
degree of oxidation desired can be accomplished, nameless
oxidation of organics can be achieved at relatively low tempera-
tures of 300°F to 400°F with this technique, compared with the
1,500°F to 2,700°F necessary for conventional incineration. Air
pollution is minimized since oxidation takes place in water at
low temperatures, and no fly ash, dust, sulfur dioxide, or
nitrogen oxide by-products are formed (1123).
Incineration has been a popular sludge treatment and
disposal method for over 20 yr. Proper treatment of sludge
prior to its introduction into the incinerator unit is the key
to successful Incineration. Pretreatment may include sludge
thickening, a macerating or disintegrating system, and/or a
dewaterlng device (1123).
The air emission products of sludge combustion are essen-
tially steam, carbon dioxide, small amounts of particulate
emissions, and oxide of sulfur (1017). Such by-products could
pose problems for human health and environmental quality.
Table 60 summarizes particulate, NOX, sulfur dioxide, and
visible emissions from four incineration sites. As shown, NOX
emissions are most significant.
165
-------
TABLE 59 . LITERATURE REVIEWED PERTAINING
TO THERMAL PROCESSES
Contami nant Reference Number
Water Quality Parameters
Ar.n.ionia 1236
BOD 864, 1435
COD 222
Nitrates 528, 788, 1281
Phosphates 507, 1354
Suspended solids 126, 1435
Other (general) 7, 88, 393, 531, 551, 552, 642, 864,
1017, 1123, 1354
Elemental Contaminants
Aluminum 507
Boron 75
Cadmium 662, 926, 1017
Chromium 436, 746, 1017
Copper 401, 746, 1017
Iron 507, 1017
Lead 662, 746, 1017
Manganese 746
Mercury 401, 926, 1017, 1077
Nickel 401, 1040
Zinc 401, 436, 1040
Biocidal Contaminants
Chlorinated
hydrocarbons 401
166
-------
TABLE 59 (continued)
Contaminant Reference Number
DDT 401
D i e1d r i n 401
Synthetic/Organic 567
Contaminants
Biological Contaminants
Bacteria 1325
Coliforms 401
Escherichia coli 1123, 1325
Mycobacterium 1123, 1325
Parasitic worms 791, 1325
Protozoa 1325
Salmonella 1123, 1325
Virus 1123, 1281, 1325
Other (general) 1123, 1281, 1325
167
-------
TABLE 60. EXHAUST EMISSIONS FROM SLUDGE INCINERATION (1017)
Emissions
Minimum
Maximum
% Carbon Dioxide -
Vol. % (dry basis)
% Excess Air (test point)
Sulfur Dioxide (ppm)
NOX (ppm)
Hydrochloric acid (ppm)
Particulate - filter (GR/SCFD)
Particulate - total (GR/SCFD)
Sludge feed to furnace
(Ib DS/hr)
Stack flow rate (SCFM)
2.2
366.0
1.97
11.47
.621
.0127
.0170
221.8
1 ,170
10.2
51.0
14.26
271.63
11.9
.0766
.0859
1 ,710.0
10,290
ELEMENTAL CONTAMINANTS
Wastewater sludges contain metals that could be hazardous
if discharged to the atmosphere during sludge incineration. The
effect of incineration on such metals will be influenced by the
forms in which the metals are found in sludge. For example, if
cadmium is present in the sludge in solution as cadmium
chloride, it will volatilize during incineration. If present
as a precipitated hydroxide, however, cadmium will probably
decompose to the oxide, but will not volatilize at the tempera-
tures of incineration. It is believed that most hazardous
metals, with the exception of mercury, will not appear dispro-
portionately in stack gases because of volatilization, but will
instead be converted to oxides and appear as particulates either
in the fly or bottom ash.
High temperatures during incineration decompose mercury
compounds to volatile mercuric oxide or metallic mercury. Tests
conducted on five incinerators equipped with scrubbers showed
an average emission factor of 1.65 g of mercury emitted to the
atmosphere per metric ton of dry sludge incinerated.
Table 61 (1017) presents concentrations of metals in the
input sludge and the bottom ash for three plants. Any discrep-
ancy in concentration would indicate volatilization of the metal
in either the fly ash or vapor.
The chemical content of ash from multiple hearth incinera-
tors reported by Olaxsey (1017) is shown in Table 62 .
168
-------
TABLE 61 . METAL TO FIXED SOLID RATIO
THREE INCINERATORS (1017)
(mg/g)
FROM
Plant A
Plant B
Plant C
Element
Sludge Ash
Sludge Ash Sludge Ash
Cadmium
Chromium
Copper
Iron
Lead
Magnesium
0.37
2.0
2.6
18.0
5.8
9.0
0.20
0.3
1.3
8.9
0.7
n.d.
n.d.
2.9
2.5
12.0
7.0
20.0
0.58
0.5
1.7
11.0
0.85
n.d.
0.31
0.7
1.6
50.0
2.0
6.0
0.22
0.6
1.6
43.0
1 .0
n.d.
TABLE 62. CHEMICAL CONTENT OF
SLUDGE INCINERATION ASH (1017)
Content
Silica (Si02)
Alumina (Alp03)
Iron Oxide (Fe203)
Magnesium Oxide (MgO)
Total Calcium (CaO)
Available Calcium
Phosphorus Pentoxide
Loss of Ignition
Percent of Total
25-30
10-13
9-10
2-2.8
30-37
1-2
7-10
0.5-1.0
Chemical analysis of wastewater sludge ashes was also
performed by Gray and Penessis (507). In general, the ashes
were composed of predominantly silt-sized particles. All the
ashes showed a pH in the range of 10.7 to 12.8 and were not
appreciably soluble in water. The soluble salt content and
conductivity of the ash leachates was almost exclusively com-
posed of dissolved calcium and magnesium. Contrary to results
on the two preceding tables, the amount of soluble iron,
potassium, and phosphate was negligible, with the leachate
containing between 0 and 0.5 mg/100 g of air-dried ash.
BIOLOGICAL PATHOGENS
Table 63 shows the effect of various pasteurization
temperatures and times on the survival of typical pathogenic
organisms found in sludge.
169
-------
TABLE 63. EFFECT OF TIME AND TEMPERATURE ON THE
SURVIVAL OF TYPICAL PATHOGENS FOUND IN SLUDGE* (1123)
Temperature C
Organism 50 55 60 65 70
-- minutes --
Cysts of Entamoeba histolytica 5
Eggs of ascaris 1umbricoides 60 7
Brucella abortus 60 3
Corynebacterium diphtheria 45 4
Salmonella typhosa 30 4
Escherichiacoli 60 5
Micrococcus pyrogene var.
aursus 20
Mycobacteriurn tuberculosis
var. promixis 20
Viruses 25
*Pathogens completely eliminated at indicated time and
temperature.
170
-------
ENVIRONMENTAL PATHWAYS
FRESH SURFACE WATER
INTRODUCTION
Direct discharge of treated wastewater to fresh surface
water is the most popular method of wastewater disposal and
potentially the most significant pathway of wastewater contami-
nants through the biosphere. In addition, relatively minor
quantities of wastewater contaminants may indirectly reach fresh
surface waters through runoff or percolation from land disposal
of wastewater effluents and sludges. This section of the report
discusses current knowledge about the fate of various contami-
nants in fresh-water systems.
Surface waters receive contaminants from treated municipal
wastewaters (as previously stated) and from other significant
sources such as agricultural runoff, storm-water discharges,
mine drainage, and atmospheric fallout. Most of the literature
pertinent to the behavior of various contaminants in fresh-water
systems - listed in Table 64 - does not differentiate contami-
nants by source. This, however, does not affect the validity of
the literature research.
Much of the material contained in this chapter was derived
from the following references: 304, 403, 464, 636, 1126, 1266,
and 1339.
WATER QUALITY PARAMETERS
General water quality parameters in surface waters are not
of direct public health concern, although they often degrade the
quality of the aquatic habitat. For example, phosphorus concen-
trations resulting from sewage disposal contributes to luxuriant
growths of certain algae, such as Anabaena, Nodularia, or NostoC'
all of which produce toxins that can be harmful to humans (1331,
1472). However, the tastes and odor also associated with such
water degradation would generally make the water unpotable long
before the concentration of toxins reached levels harmful to
public health.
Suspended solids in wastewater can carry adsorbed viral and
other biological contaminants (1476). Trace metals occur in
higher concentrations when associated with suspended matter than
when they are in a dissolved state (227), a phenomenon that will
be discussed more fully in the following section on elemental
171
-------
TABLE 64. LITERATURE REVIEWED PERTAINING TO
WASTEWATER TREATMENT PLANT EFFLUENT
DISPOSAL TO FRESH-WATER SYSTEMS
Contaminant
Water Quality Parameters
Reference Number
Ammonia
BOD
COD
Chlorides
Cyani des
Fluorides
Ni trates
Ni tri tes
Oil and grease
Phosphates
46, 47, 54, 133, 138, 221, 364, 418, 454,
757, 829, 920, 1002, 1125, 1156, 1239, 1306,
1407, 1436, 1453
46, 133, 354, 364, 533, 757, 882, 1064,
1283, 1306, 1392, 1405, 1470
46, 108, 133, 418, 1156, 1283,1306, 1470, 1475
46, 47, 54, 133, 500, 1018, 1156, 1177,
1306, 1405, 1470
1306, 1436
54, 133
9, 46, 47, 138, 151, 190, 221, 364, 411,
443, 454, 500, 533, 572, 600, 663, 757,
793, 829, 883, 1000, 1002, 1027, 1076,
1125, 1156, 1239, 1306, 1331, 1407, 1453
46, 133, 138, 221, 454, 663, 757, 793,
829, 920, 1125, 1156, 1239, 1407
757, 1306, 1497
46, 47, 54, 133, 138, 190, 221, 364, 418,
454, 533, 555, 600, 663, 757, 784, 829,
882, 883, 1156, 1306, 1314, 1331, 1407,
1436, 1448, 1472
Suspended solids 47, 227, 354, 454, 757, 920, 1006, 1125,
1171, 1306, 1470, 1476
Total dissolved
solids
Total organic
carbon
364, 418, 454, 757, 1125, 1177, 1306
133, 1125, 1167, 1264, 1283, 1306
Other (general) 304, 336, 385, 403, 464, 757, 805, 1125,
1126, 1156, 1253, 1258, 1266, 1286, 1370
172
-------
TABLE 64. (continued)
Contaminant
Elemental Contaminants
Aluminum
Reference Number
Antimony
Arsenic
Barium
Beryl 1ium
Boron
Cadmi urn
Chromium
Cobalt
Copper
Germanium
Iron
28, 634, 636, 968, 1082, 1093, 1121,
1125, 1266, 1339, 1380, 1407, 1453
303, 636, 777, 1093, 1121, 1266, 1306,
1339, 1380, 1453
303, 360, 409, 418, 520, 636, 1093,
1121, 1138, 1266, 1306, 1339, 1380,
1453
23, 418, 636, 968, 1093, 1121, 1266,
1339, 1380, 1453
418, 636, 1093, 1121, 1266, 1339, 1380,
1453
28, 133, 418, 636, 757, 1093, 1113, 1121,
1266, 1339, 1380, 1453
77, 117, 151, 175, 190, 263, 303, 360,
418, 433, 441, 462, 524, 629, 636, 724,
740, 757, 859, 1075, 1082, 1093, 1121,
1124, 1125, 1171, 1195, 1266, 1306, 1355,
1380, 1407, 1433, 1453, 1523, 1526, 1538
151, 263, 365, 418, 473, 629, 777, 859,
1082, 1093, 1121, 1125, 1128, 1306, 1380,
1407, 1433, 1436, 1526, 1538
303, 445, 473, 600, 634, 636, 740, 782,
859, 968, 1056, 1074, 1075, 1082, 1093,
1121, 1125, 1266, 1339, 1380, 1456, 1538
28, 77, 151, 190, 206, 418, 433, 445, 451,
473, 504, 505, 629, 634, 740, 777, 782,
859, 968, 1044, 1074, 1075, 1082, 1093,
1121, 1125, 1171, 1266, 1339, 1355, 1407,
1433, 1453, 1454, 1543, 1562
532, 636, 1093, 1121, 1266, 1339, 1453
28, 263, 306, 438, 473, 528, 634, 636,
782, 852, 854, 968, 1074, 1075, 1093,
1121, 1125, 1266, 1339, 1355, 1371, 1380,
1407, 1562
173
-------
TABLE 64. (continued)
Contaminant Reference Number
Lead 117, 151, 190, 204, 206, 360, 418, 441, 445,
462, 511, 524, 629, 634, 636, 724, 740, 757,
777, 782, 859, 912, 968, 1018, 1060, 1074,
1075, 1082, 1088, 1093, 1121, 1171, 1174,
1306, 1339, 1355, 1380, 1391, 1433, 1453,
1508, 1544, 1562
Manganese 28, 190, 206, 263, 306, 418, 445, 462, 473,
600, 636, 645, 844, 1074, 1075, 1121, 1171,
1266, 1306, 1339, 1355, 1380, 1433, 1453,
1523, 1538, 1544
Mercury 12, 37, 40, 78, 110, 117, 151, 269, 303,
418, 455, 462, 512, 524, 591, 607, 630,
634, 636, 726, 741, 752, 757, 760, 777, 809,
861, 968, 988, 1082, 1089, 1093, 1121, 1125,
1266, 1306, 1339, 1355, 1359, 1378, 1379,,
1380, 1407, 1433, 1446, 1453, 1526, 1543, 1562
Molybdenum 303, 636, 1093, 1121, 1266, 1339, 1380, 1453
Nickel 28, 263, 418, 436, 445, 473, 629, 636, 782,
859, 968, 1074, 1093, 1121, 1125, 1171, 1266,
1306, 1355, 1380, 1407, 1453, 1544
Selenium 418, 636, 1093, 1114, 1121, 1266, 1306, 1339,
1380, 1453
Thorium 636, 1093, 1266, 1339, 1380, 1453
Tin 636, 1082, 1093, 1266, 1339, 1453, 1380
Uranium 636, 1093, 1121, 1266, 1339, 1380, 1453
Zinc 21, 77, 151, 190, 263, 303, 418, 441, 445,
504, 505, 524, 629, 636, 724, 740, 782, 844,
859, 968, 1074, 1075, 1082, 1093, 1121, 1125,
1171, 1266, 1306, 1319, 1339, 1355, 1380,
1407, 1433, 1453, 1523, 1538, 1544, 1562
Other (general) 133, 228, 636, 1211, 1339, 1380, 1454, 1508
Biocidal Contaminants
Aldrin 139, 304, 373, 403, 783, 846, 1150, 1237, 1418
174
-------
TABLE 64. (continued)
Contaminants
Arsenated 1237
hydrocarbons
Chlorinated 139, 275, 304, 324, 373, 387, 403,
hydrocarbons 418, 464, 501, 512, 577, 757, 759, 772,
778, 827, 846, 998, 999, 1022, 1023,
1049, 1068, 1125, 1132, 1170, 1220,
1237, 1532, 1543
ODD 373, 403, 429, 577, 757, 759, 846,
1022, 1225
DDE 139, 373, 403, 429, 464, 529, 577, 759,
846, 972, 1125, 1150
DDT 14, 139, 304, 325, 373, 403, 420, 429,
453, 464, 515, 577, 641, 706, 757, 759,
846, 902, 931, 972, 1022, 1023, 1125,
1150, 1170, 1180, 1220, 1306, 1418, 1532
Dieldrin 373, 403, 429, 846, 972, 1125, 1550,
1237, 1306, 1418, 1470
Endrin 373, 403, 464, 846, 1150, 1418
Herbicides 597, 627, 1125, 1220, 1488, 1551
Organophosphorus 304, 306, 373, 418, 451, 501, 566, 846,
pesticides 1196, 1245, 1550
Soil sterilants 73
Other (general) 291, 304, 418, 478, 600, 641, 827, 846,
846, 1049, 1090, 1225, 1550, 1454, 1512
Synthetic/Organic 58, 149, 166, 275, 587, 770, 808, 856,
Contaminants 879, 998, 999, 1029, 1126, 1323, 1331,
1340, 1420, 1460, 1558
Biological Contaminants
Adeno virus 757, 1509
Bacteria 74, 140, 197, 231, 243, 296, 304, 418,
434, 521, 588, 596, 656, 676, 678, 786,
815, 938, 1131, 1202, 1306, 1370, 1399
Clostridium 1131
botulinium
Clostridium welchi 1131
175
-------
TABLE 64. (continued)
Contaminants Reference Number
Coliforms 46, 133, 140, 147, 187, 188, 197, 248,
299, 307, 354, 418, 457, 596, 612, 655,
656, 663, 757, 938, 1064, 1131, 1202,
1263, 1268, 1275, 1283, 1366, 1436,
1458, 1475, 1509, 1518
Coxsackie virus 11, 588, 757, 1366, 1509
(A+B)
ECHO virus 11 , 588, 757, 1366
Escherichia coli 147, 180, 588, 655, 1275
Fecal
streptococci 140, 147, 180, 197, 307, 457, 458, 612,
656, 1131 , 1202, 1263, 1283
Hepatitis virus 163, 257, 380, 492, 890, 1366, 1509
Leptospirosis 380
Polio virus 11, 85, 492, 750, 1209, 1366, 1509
Protozoa 655
Salmonella 30, 85, 231, 248, 307, 457, 588, 671,
1283
Shigella 30, 85, 588, 1449
Staphylococcus 1131
aureus
Vibrio cholerae 74, 1331
Virus 46, 94, 99, 111, 133, 246, 251, 257,
380, 454, 468, 564, 588, 678, 750, 757,
815, 826, 877, 890, 938, 1182, 1207,
1237, 1240, 1283, 1331 , 1366, 1434,
1443, 1478
Yeasts 596
Other (general) 180, 250, 433, 786, 1225, 1370
176
-------
contaminants. Suspended matter and dissolved organics from
wastewater can affect the natural biological community which,
in turn, affects the nitrogen interconversions and nitrate con-
centrations.
The nitrogen-containing compounds (ammonia, nitrite, and
nitrate) are theoretically hazardous because of nitrate's
association with methemoglobinemia. Ammonia and nitrite nitrogen
can be readily converted into nitrate by chemical or biological
reactions. There are, however, no reported cases of detrimental
public health effects resulting from nitrates in surface waters.
This is because the dilution and natural processes occurring in
surface waters prevent nitrate concentrations from reaching
health impairing levels. There are, however, several reported
cases of groundwater nitrate contamination rising to dangerous
levels. These cases are discussed in the land/groundwater
section of this report.
ELEMENTAL CONTAMINANTS
The behavior of elemental contaminants in fresh-water
systems is very complex. Generally, elemental contaminant
transport mechanisms can be divided into either (1) elements in
solution, or (2) elements associated with inorganic or biologi-
cal particulates. Each of these mechanisms can be broken down
still farther. Dissolved elements may occur as unassociated
ions or as inorganic or organic complexes. Elemental contaminant/
inorganic particulate associations include coulombic attraction,
as in conventional adsorption; ionic bonding, as in ion exchange;
precipitated or coprecipitated metal coating; or incorporation
into particulate crystalline lattices. Elemental contaminant/
biological particulate associations include surface adsorption,
ingested particulation, and biochemical incorporation into the
organism. The particular transport mechanism that will predomi-
nate in a given water system depends, in part, on the geohy-
drologic environment, mineralogy/petrology of the river or lake
bed, pH, temperature, dissolved organic or oxygen content,
biological activity, elemental type and source, and nonelemental
chemical composition of the water.
This variety of factors does much to explain the seeming
discrepancies in the work of different researchers attempting
to establish elemental distributions in fresh-water systems.
For instance, Gibbs (473), in his examination of the Yukon and
Amazon Rivers, concluded that precipitated metal coating and
crystalline incorporation accounted for approximately 90 percent
of the transported iron, nickel, copper, chromium, cobalt, and
manganese (Table 65 ). Perhac (1074), on the other hand, in
his analysis of two Tennessee streams, concluded that 95 percent
of the total stream content of cadmium, cobalt, copper, nickel,
lead, and zinc was in the dissolved state (Table 66). Assuming
that there was no gross experimental error, widely differing
environmental factors must have prevailed.
177
-------
TABLE 65 . PERCENTAGES OF THE TOTAL AMOUNTS OF
IRON, NICKEL, COBALT, CHROMIUM, COPPER, AND MANGANESE
TRANSPORTED BY FIVE MECHANISMS IN THE
YUKON AND AMAZON RIVERS (473)
Mechanism
In solution and
organic complexes
Adsorbed
Precipitated and
copreci pita ted
In organic solids
In crystalline
sediments
In solution and
organic complexes
Adsorbed
Precipitated and
copreci pi tated
In organic solids
Iron Nickel
Amazon River
0:7 2.7
0.02 2.7
47.2 44.1
6.5 12.7
45.5 37.7
Yukon River
0.05 2.2
0.01 3.1
40.6 47.8
11.0 16.0
Cobalt
1.6
8.0
27.3
19.3
43.9
1.7
4.7
29.2
12.9
Chromium
10.4
3.5
2.9
7.6
75.6
12.6
2.3
7.2
13.2
Copper
6.9
4.9
8.1
5.8
74.3
3.3
2.3
3.8
3.3
Manganese
17.3
0.7
50
4.7
27.2
10.1
0.5
45.7
6.6
In crystalline
sediments 48.2 31.0 51.4 64.5 87.3 37.1
178
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TABLE 66. HEAVY METAL DISTRIBUTION IN STREAMS (1074)
Average percentage of element occurring in specified form
Dissolved Coarse
Solid Participate Colloid
Cadmium 92.9 5.1 2.0
Cobalt 91.8 7.8 0.4
Copper 93.2 5.4 1.4
Iron 19.6 77.0 3.3
Lead 90.8 8.0 1.2
Manganese 23.4 74.6 2.0
Nickel 93.2 6.3 0.5
Zinc 77.8 21.8 0.4
179
-------
The complexity of the chef.iistry, biology, and physics
involved in water behavior of elements precludes a detailed dis-
cussion. Instead, a brief discussion is presented of the more
important aspects.of the behavior of elements in water, followed
by a detailed examination of a few sample elements (mercury,
arsenic, lead, cadmium, copper, iron) to demonstrate the princi-
ples involved.
Dissolved elements may occur as unassociated ions or as
inorganic or organic complexes. Of the major elements under
discussion in this report, only barium appears to any great
extent as the unassociated cation; barium ions do not hydrolyze
and form only weak complexes. However, several of the elements
occur as unassociated anions: antimony, arsenic, boron, chromium,
molybdenum, and selenium generally occur in fresh-water systems
as the-oxo anion. This is largely because the major sources of
these elements, including wastewater, are rich in the anionic
forms, and because the cationic forms are easily oxidized to the
oxo anion in aquatic systems.
The majority of the dissolved elements normally exist as
inorganic or organic complexes. Table 67 lists the more common
ligands and the conditions and elements normally associated
with them. In relatively pure water, aquo (1^0) or hydroxo
(OH") complexes are formed. At pH levels above neutral, many of
the metal-hydroxo complexes are converted to metal hydroxides or
oxides, which will precipitate out of solution or behave as
colloids.
The other inorganic ligands responsible for keeping metals
in solution in natural waters include carbonate, halides
(notably chloride and fluoride), sulfur species (SH~, sulfate,
and sulfite), and nitrogen species (ammonia, nitrate, and
nitrite). Most of the complexes formed from these ligands are
thermodynamically unstable and appear as transition states
between the free metal ion and a precipitate. The complexes,
however, serve to keep the metals in solution for a time, and
play a role in dissolving otherwise insoluble metals from
precipitates or crystalline lattices.
Of somewhat more importance in terms of complex stability
are the multidentate organic ligands. One of the reasons for
the lack of inorganic complex stability is that most of the
inorganic ligands are monodentate, i.e., there is one ligand for
each metal ion coordination site. Organic ligands are multi-
dentate, i.e., a given ligand can usually bond to two or more
of a given ion's coordination sites. A multidentate complex
(chelate) is more stable than a corresponding complex with
monodentate ligands; thus, a chelate complex is apt to keep a
metal ion in solution far longer than will an inorganic complex.
Normally, this is not a problem; in relatively unpolluted fresh
water the organic content is low, but in highly polluted water
180
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TABLE 67. (continued)
Footnotes
5
The solubility falls markedly in the presence of this ligand
at above pH due to precipitation of a carbonate or similar
basic compound.
Coordination occurs only at pH above 7 due to ligand insta-
bility, etc.
Coordination occurs only at pH above 8-9.
pPrecipitation almost always occurs. If nothing is marked,
there is no coordination of this metal by this ligand in
natural waters.
aWater will only coordinate if no other stronger ligand is
present. In some cases, there is an equilibrium.
Bromide and iodide resemble chloride except that they both
precipitate silver, whereas silver chloride is fairly
soluble due to AgCl?-ions at high chloride concentrations.
Iodide also precipitates copper and gold.
cTwo valent iron in absence of air only.
If ammonia is absent, a complex may be formed.
eBicarbonate usually forms carbonate complexes, but metals
so marked have a soluble bicarbonate which is water co-
ordinated. Be and Tl have soluble water coordinated car-
bonates, and Ag has both sparingly soluble water coordi-
nated carbonate and hydroxide.
182
-------
unusually high soluble metal concentrations may result. Further-
more, there is evidence that some synthetic organic ligands,
such as may be found in wastewater, form stronger complexes than
do natural organic ligands (782). It has been demonstrated that
one synthetic ligand, nitri1otriacetate (a proposed substitute
for phosphates in detergents), is capable of dissolving signifi- •
cant quantities of precipitated lead out of bottom sediments
(511, 1562).
The tendency for a given metal ion and ligand to form
complexes depends on solution pH, concentration of the metal
and ligand, concentrations of other metals and ligands in
solution, equilibrium constants, redox conditions, and so forth.
No two elemental contaminants behave exactly alike, and even
different oxidation states of the same element may exhibit
widely varying solution chemistry. Moreover, natural water
systems are seldom in an equilibrium condition before, much less
after, wastewater addition. This constantly changing system
makes concise pathways almost impossible to construct. In
general, low pH, an oxidizing environment, and the presence of
a variety of ligands enhance solution tendencies. This is signi-
ficant from a public health standpoint, because soluble element
species are much more readily available for human contact than
are precipitates or particulate elements.
Elemental contaminant/inorganic particulate interactions
typically account for the bulk of the nondissolved element, fraction
These interactions include coulombic or ionic attraction,
precipitated or coprecipitated coating, or lattice incorporation.
Coulombic attraction, or adsorption, is the least important
transport mechanism except in the case of colloidal particulates,
such as microparticulate iron or manganese hydroxides, which
carry a weak negative charge and attract elemental cations.
Larger particulates do not possess a strong enough charge to
make coulombic attraction important.
Ionic attraction, or ion exchange, is somewhat more impor-
tant. In this process, the heavy metals (Mn, Fe, Co, Ni , Cu,
Zn, Mo, Cd, etc.) replace alkali and alkali earth cations (K,
Na, Li, Mg, Ca) attached to crystalline lattices by ionic bonds.
These ionic bonds will hold unless (1) the element is displaced
by another element forming a stronger ionic bond, (2) a ligand
forming a coordination bond stronger than the ionic bond breaks
the ionic bond, or (3) an excess of alkali earth cations is
available to force the equilibrium back to its original state.
According to Gibbs (473), lattice incorporation is generally
insignificant as a transport or removal mechanism of dissolved
wastewater elemental contaminants. It is a slow process that
takes place in the sediments and has a more significant impact
on the elemental composition of the sediments. The lattice-
incorporated elemental burden of a water system is primari lyfrom
weathered rock, sand, and clay.
183
-------
Precipitated and coprecipitated metal coatings account for
most of the non-native participate element content of a water
system. Under favorable pH-EH conditions, metal precipitates
will form. The initially small precipitate particles will tend
to agglomerate or adhere to any available surface. In addition,
cations held to participates by coulombic or ionic bonds can
form covalent bonds with anionic components of the particulate,
if solution redox potentials change. In either case, the
particulates are left with a coating of metal precipitates.
This coating ultimately settles out of solution with the partic-
ulate, removing the elements from possible ready human
contact unless they are redissolved by a change in redox
conditions, a strong ligand, or some other agent strong enough
to attach the precipitate. These coatings may account for the
bulk of the inorganic particulate element burden of waste-
waters .
Elemental contaminants incorporated with biological particu-
lates constitute the remainder of the particulate elemental
burden of a water body or wastewater. This incorporation may
take the form of surface adsorption, inorganic particulate
ingestion, or biochemical incorporation into an organism's
tissues. Obviously, the role played by biological transport is
highly dependent on the 'ype and quantity of organisms present.
Surface adsorption is usually associated with microorganisms.
Particulate ingestion is associated with those organisms that
have internal digestive organs (e.g., fish, crustaceans, worms,
etc.); ingested particulates are usually eliminated within a
short time, but the elements associated with them may be bio-
chemically incorporated into the organism's tissues. From a
public health standpoint, soluble and biochemically incorporated
elemental contaminants are the most important, for it is by
these routes that potentially hazardous elements reach man.
Biochemical incorporation involves both essential trace element
concentration (e.g., cobalt in vitamin B-12) and reaction of an
element with cellular chemicals (e.g., the reaction of mercury
with sulfur-containing amino acids in proteins). Both plants
and animals are involved and, thus, the concentration of a
given element may move up the food chain. The incorporation
is reversible; once the organism in its environment is removed
from contact with the element, the latter can gradually be
excreted .
No two elements behave exactly alike; furthermore, the
number of factors available that can affect transport mechanisms
makes the possibilities nearly endless. There are similarities,
however, that make the use of examples illustrative and useful;
mercury, arsenic, iron, cadmium, copper, and lead will bi used
to provide detailed descriptions of the general pathways
disucssed above.
184
-------
Mercury from wastewater enters a water system primarily as
the metal or divalent cation. Although of limited solubility,
it can reach concentrations of 100 ppb in aerated water (455).
Metallic mercury alone is soluble up to 25 ppb and will hydrolize
to soluble Hg (OH)2 in oxygenated systems, increasing the overall
solubility and water content. Despite these solubility figures,
mercury concentrations seldom exceed 5 ppb (Table 68 } except
in polluted water.
Inorganic and biological adsorption, absorption, and pr'e-
cipitation serve to keep the concentrations of dissolved mercury
much lower than the theoretical maximum. In general, the bulk
of the mercury in a given water system is in the sediments;
Table 69 gives a summary of some of the factors affecting
mercury incorporation in the sediments. In reducing sediments,
mercury i.s tied up as the sulfide, although if the system be-
comes sufficiently alkaline, HgSp may be released into solution.
Should the sediments become aerobic, the sulfide will be oxi-
dized to sulfate, and the mercury will be released.
All soluble mercury species except mercuric sulfide can be
absorbed by bacteria. Once the mercury is in the bacteria, a
series of transformations - possibly via a detoxification
mechanism - convert the incorporated mercury into mono- and
dimethyl mercury, both soluble at low concentrations and readily
released into solution. The methyl mercury compounds are much
more 1 ipid-preferring than the inorganic forms and are quickly
absorbed by living tissues. As a rule, mercury concentrations
tend to increase in organisms up the food chain, so that the
highest concer.trat ions are found in fish. This is partly due to
absorption of methyl mercury from the water and partly from
ingestion of plants or smaller organisms containing methyl
mercury. When the organisms die, the mercury returns to the
sediments, where most of the bacterial methylation occurs.
Table 70 lists some sample sediment and plankton/algae mercury
concentrations in Lake Erie.
Arsenic, selenium, and antimony are chemically similar and
exhibit analogous environmental behavior. Arsenic has been
studied far more than either selenium or antimony. The following
discussion of arsenic is largely applicable to selenium and
antimony as we!1.
Arsenic has an unusually complex chemistry in aquatic
systems: oxidation-reduction, ligand exchange, precipitation,
adsorption, and biomethylation reactions all take place.
Arsenic species can be removed from water via surface adsorption
and coprecipitation with metal ions; both arsenate (As04~3) and
arsenite (As03~3) have a high affinity for hydrous iron oxides
and readily coprecipitate with or adsorb onto them. Signifi-
cantly, iron ores are always enriched with arsenic (409).
Aluminum hydroxide and clays are adsorb arsenate species,
although to a lesser degree.
185
-------
TABLE 68 . SELECTED CONCENTRATIONS OF
MERCURY IN NATURAL WATERS (455)
Source and Location Mercury (ppb)
River water, European USSR 0.4-2.8
River water, Armenia 1-3
Saale River, Germany 0.035-0.145
River water, Italy 0.01-0.05
River water, near mercury deposits, Italy up to 136
Colorado River, Arizona <0.1
Ohio River, II1inois 0.1
Mississippi River, Kentucky <0.1
Missouri River, Montana <0.1
Missouri River, St. Louis, Missouri 2.8
Kansas River, Topeka, Kansas 3.5
Hudson River, New York 0.1
Lake Champlain, New York <0.1
Maumee River, Antwerp, Ohio 6.0
Delaware River, New York <0.1
186
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187
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TABLE 70. MERCURY CONTENT OF SEDIMENTS AND PLANKTON/ALGAE SAMPLES
COLLECTED FROM LAKE ERIE (1089)
Station
No.
01
02
03
04
05
06
07
08
09
10
11
12
13
14
15
16
Approximate
Location 5
Buffalo River
Cattaraugus Creek
Barcelona
Ashtabula
Fairport
Cleveland
Toledo
Detroit River
Mid. Bass Island
Port Crewe
Port Stanley
Long Point
Long Point Bay
Port Mai tl and
Mid-Lake
Black Rock Channel
Mercury content in yg/g*
Sediments** Plankton/Algae
2.0 31.2
1.2 25.1
0.6 2.8
4.6 7.4
1.5 12.8
12.0 33.5
10.4 20.5
4.5 26.1
1.5 20.1
0.5 12.4
1.5 12.0
7.0 14.7
1.0 23.7
1.8 15.4
1.5 0.6
12.4 27.8
**
In terms of the equivalent dry wt of the sample.
Sediment samples from 3 to 30 cm below the water-sediment interface.
188
-------
Microbial transformations of arsenic, while demonstrable
in the laboratory, have not been positively identified in
natural water systems. The two most commonly postulated trans-
formations are oxidation of arsenite and methylation. Methyla-
tion is important because it could be a means by which sediment
arsenic is recycled back into the water system; natural aquatic
methylation has not been demonstrated.
Soluble iron entering a water system in wastewater will
usually be either ferrous (Fe II) or complexed ferric (Fe III)
iron. The former is much more soluble than the latter (which
has a stronger tendency to form complexes), although neither
tends to remain in solution long. In the surface layers of
most natural waters, pH levels and oxygen conditions are such
that Fe (II) is readily oxidized to Fe (III), which just as
readily hydrolyzes to insoluble hydrous ferric oxide (FeOOH).
Hydroxide has a much stronger affinity for ferric iron than do
basic organic or inorganic ligands.
Hydrous ferric oxide tends to form microcrystalline preci-
pitates of a colloidal nature, so that it is almost impossible
to analytically distinguish between soluble and colloidal iron.
Consequently, the two forms of iron are usually reported together
as soluble iron. Although hydroxide supersedes other an ionic
ligands, frequent incorporation of coordinating anions into
ferric oxide precipitates enhances colloidal stability and
further blurs the distinction between the colloid and soluble
ferric complexes.
Cadmium readily precipitates as the hydroxide or carbonate
and consequently is not normally found in high concentrations
in surface waters. In fact, several researchers (360, 968) have
noted that high soluble cadmium concentrations are invariably
associated with polluted water that receives a steady cadmium
source, such as industrial wastewater.
Cadmium (II) readily hydrolyzes and forms transitory
inorganic complexes, such as chloride complexes that have a
limited affinity for hydrous iron and manganese oxides, and
organic particulates. The organic affinity probably indicates
a reaction between the cadmium and sulfur-containing compounds.
Cadmium forms the insoluble hydroxide at pH levels of 7
and above; it forms the insoluble carbonate under oxidizing
conditions, particularly in soft waters where cadmium does not
have to compete with calcium and magnesium for the carbonate
an ion. Once cadmium has precipitated and settled into the
sediments, it is not readily removed. Consequently, if cadmium
additions are reduced, a water body will tend to purify itself
of soluble cadmium.
Copper, and to a lesser extent nickel, occupy an unusual
position in water chemistry and biology because they are both
189
-------
nutrients and toxins. This has a pronounced effect on their
water chemistry. Copper contained in wastewater may be either
soluble or particulate; neither form predominates as a rule.
Copper adsorbs readily onto clay and organic particulates.
Copper also forms several very stable complexes. In pure water,
while the aquo complex may predominate, the carbonate, chloride,
and amine inorganic complexes are much more stable.
Ultimately, the stability soluble of copper can be
attributed to organic complexes, since copper forms coordination
complexes with virtually every conceivable organic ligand.
These complexes are very stable thermodynamically and are also
resistant to microbial attack, a mechanism responsible for the
destruction of most organic complexes. Copper is a bacterial
toxin and, if released from its complex by microbial attack,
simply kills the offending bacteria and forms a new complex (968).
Copper is removed from solution via precipitation or
biological incorporation. Since the most common precipitate is
the carbonate, most sediment copper is in the carbonate form
(263, 968). An essential trace nutrient, copper is readily
incorporated into aquatic plants and animals.
The main soluble species of lead in wastewater are the
lead (II) cation and the hydrolyzed complex Pb (OH)3~ . Lead
forms a variety of stable complexes as well; researchers (777)
have identified both PbOH+ and Pb(C03)2~2 in natural water
systems. Lead complexes easily with a variety of organic
chelates,. forming very stable complexes. Some of these com-
plexes are more stable than the precipitated lead in the sedi-
ments; therefore, they will actually dissolve otherwise in-
soluble lead. A case in point is nitrilotriacetate, which
can solubilize lead from lead carbonate precipitates (511).
Low water organic content generally prevents solution lead
concentrations from exceeding a few parts per billion. In most
water systems, lead introduced with wastewater readily forms
insoluble Pb(OH)2 and PbCC^, which will precipitate and adsorb
onto suspended particulates. Ionic lead is not so strongly
adsorbed, although it does have some affinity for clays.
Hydrous iron oxides strongly sorb ionic lead at neutral
to slightly acidic pH levels. Some ionic or complexed lead
adsorbs onto or is chelated with the surface mucilage of algae,
and microorganisms immobilize substantial quantities of inorganic
lead, presumably on or in all membranes (1391). As a result of
all of these adsorption/precipitation mechanisms, most of the
water lead burden is associated with particulate matter, and
most of the lead entering a normal water system ultimately
finds its way into the sediment.
190
-------
Natural water bodies normally contain very low dissolved
concentrations of the more harmful elemental contaminants.
Unless wastewater additions are voluminous and repeated, natural
water chemistry can purify the water of soluble species fairly
well. However, there is a buildup of these elementals in the
sediments. This means that if the local water chemistry should
change significantly, the elements can still be released to
solution. Natural water purification mechanisms can only change
the wastewater elemental problem from a real to a potential
hazard; they cannot solve the problem of elemental contamination.
BIOCIDAL CONTAMINANTS
The single most important source of biocidal contaminants
in fresh-water bodies is surface runoff, followed by aerial
fallout and industrial waste discharge from plants manufacturing
biocides. In general, municipal wastewater will have detectable
quantities of biocides only if it contains biocide manufacturing
wastes. Cleanup and disposal by households, farmers, gardeners,
etc., contribute minimally to the overall wastewater burden.
In this discussion, biocides will include chlorinates,
hydrocarbons, organophosphates, carbamates, and ionic bio-
cides (Table 71 ) .
Biocides can be transported or removed from a system by
microbial or chemical degradation, photodegradation, sediment
or humic matter, adsorption, volatilization, and biological
uptake. All of these mechanisms are in turn affected by pH,
temperature, salt or organic content, and bioproductivity. One
mechanism that has not been studied to any great degree -
especially in fresh water - is aerosolization. This mechanism
will be discussed in greater detail in the marine water section
of this chapter, as its importance in fresh-water systems is
limited mainly to the larger lakes. Briefly, however, aerosoli-
zation occurs through the action of wind and waves on floatables.
Aerosols and particulates can be released to the air and trans-
ported great distances by the wind; biocides can concentrate in
floatables via dissolution in surface oil, adsorption on floating
matter, or flotation (caused by low specific gravity and insolu-
bility), thus becoming amenable to the aerosolization process.
As mentioned above, this is not an important transport mechanism
in most fresh-water systems.
Various transport mechanisms affect the biocide classes
differently. This is demonstrated in Table 72 , which compares
the persistence of selected chlorinated hydrocarbons, organo-
phosphates, and carbamates in river water Because of these
differences, the biocide classes will be discussed seoarately.
191
-------
TABLE 71 . BIOCIDE TYPES AND EXAMPLES
Chlorinated Hydrocarbons
DDT (ODD, DDE)
Methoxychlor
Endrin
D i e 1 d r i n
Aldrin
Toxaphene
Llndane
Chlordane
Heptachlor
Organophosphates
Parathion
Malathlon
Dlmethoate
Methyl parathlon
Phorate
Demeton
Ethlon
Dlsulfaton
Carbamates
Carbaryl
S e v i n
Baygon
Pyrolan
D i m e t i 1 a n
Ionic BiPC ides
Diquat
Paraquat
Chlormequat
Morfaunquat
Phosphon
Hyami ne
2,4-D
2,4, 5-T
Dalapon
Silvex
Dichlobenil
192
-------
TABLE 72. PERSISTENCE OF COMPOUNDS IN RIVER WATER (846)
Compound
Organochlorine compounds
BHC
Heptachlor
Adlrin
Heptachlor epoxide
Telodrin
Endosulfan
Dieldrin
DDE
DDT
ODD
Chlordane (tech.)
Endrin
Organophosphorus compounds
Parathion
Methyl parathion
Malathion
Ethion
Trithion
Fenthlon
Dimethoate
Merphos
Merphos recov. as
Azodrin
Original compound found, percent
0-time
2 wk
4wk
~8wk
Def
Carbamate compounds
Sevin
Zectran
Matacil
Mesurol
Baygon
Monuron
Fenuron
100
100
100
100
100
100
100
100
100
100
100
100
100
25
100
100
25
30
100
100
100
100
90
100
100
0
80
100
10
5
100
100
100
100
85
100
100
0
40
100
0
0
100
100
100
100
85
100
100
0
20
100
0
0
100
100
100
100
85
100
100
80
100
100
90
100
100
0
100
100
50
25
25
90
25
50
100
0
50
100
30
10
10
75
10
10
85
0
30
100
<5
0
0
50
0
0
75
0
10
100
0
0
0
50
0
0
50
0
<5
100
90
100
100
90
100
80
80
5
15
60
0
50
40
60
0
0
10
0
30
30
20
0
0
0
0
10
20
0
0
0
0
0
5
0
0
193
-------
The chlorinated hydrocarbon pesticides are all insoluble
in water, with the exception of lindane, which is sparingly
soluble to iO ppm (403)- They are generally resistant to
microbial and chemical degradation, as evidenced by their
estimated environmental half-lives, shown in Table 73 .
TABLE 73 . ESTIMATED PESTICIDE HALF-LIVES (403)
Pesticide HalJr-LifgJ,j jr^s
Lindane 2
Chlordane 8
Toxaphene 11
Heptachlor 2 to 4
DDT 10 to 20
Endrin (Dieldrin) 8 to 10
These pesticides are somewhat more susceptible to photo-
degradation, although the degradation products are often as
toxic as the parent compound, regardless of the type of
degradation, DDT is decomposed chemically to ODD and DDE and
photochemically to PCB's (304, 1170); aldrin is photooxidized
to the more toxic dieldrin (304); and methoxychlor is degraded
to methoxychlor DDE (1049). Surface oil slicks tend to concen-
trate chlorinated hydrocarbons and thus make them more available
for photochemical degradation (304).
Chlorinated hydrocarbons in general readily adsorb onto
fungi, algae, and floe-forming bacteria (783, 1049), and thus
tend to concentrate in biological communities. When ingested
by higher organisms, they accumulate in lipid tissues; conse-
quently, there is a tendency for chlorinated hydrocarbons to
concentrate up the food chain.
Chlorinated hydrocarbon insecticides differ in chemical
structure, but they all exhibit affinity for organic sediments
and resistance to microbial attack. As a result, there is
accumulation in bottom sediment. Research on Lake Michigan
demonstrates this, as shown in Table 74 . Routh (1180) showed
that DDT, with its affinity for fine particulate clay sediments,
concentrated up to 20 times normal background levels, from 10
to 200 ppb. This affinity for organic matter and particulates
leads to high sediment DDT concentrations. Table 75 shows DDT
concentrations in stream sediments over a period of time.
Adsorption of DDT on algae can be 10 to 100 times greater than
adsorption on clay (706). Moreover, DDT seems to have an
inhibitory effect on sediment bacteria (14).
There has been little research on other chlorinated hydro-
carbons, much of it limited to an evaluation of environmental
194
-------
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195
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TABLE 75. DDT CONCENTRATIONS IN STREAM SEDIMENTS (515)
Years after DDT
one (ppm)
application muds
0 .83
(.04-1.6)
1 1 .08
(.25-1.9)
2
3 .21
(.12-.30)
4 .21
(.16-.25)
5 .59
(.07-1.9)
6 .06
7 .21
(.11-.31)
8 .13
9 .03
(.02-.04)
10 .07
(0-.16)
Never sprayed .006
(0-.02)
Two sprays .43
(.17-.77)
Three sprays .35
(.17-.6)
196
-------
levels. Table 76 gives the results of one such survey for
d i e 1 d r i n.
TABLE 76 . DIELDRIN IN RIVER BOTTOM SILTS (972)
Source Dieldrin (ppb)
Iowa River 8.8
Des Moines River 35
East Nishnabotna River 21
West Nishnabotna River 16
Upper Iowa River <1
Johnson County Creek 170
Other researchers have found that the highest reported
concentrations of several pesticides in major U.S. river basins
from 1958 to 1965 were as follows: dieldrin, 0.100 vg/l; endrin,
0.116 vg/l; and DDT, 0.148 yg/£. Dieldrin was the most widely
found pesticide (1470).
The organophosphorus biocides are more soluble than the
chlorinated hydrocarbons. These solubilities range from 1 ppm
for ethion to 20,000 ppm for dimethoate; most fall in the 25 to
150 ppm range (403). The organophosphorus biocides are also
more amenable to both microbial and chemical degradation. Even
parathion, the most chemically resistant of the organophosphates,
will degrade via ester linkage hydrolysis in a few months under
normal conditions. The degradation takes place in just a few
weeks in polluted water with a high bacteria count (501). Yu
and Sanborn's (1550) experimental evaluation of parathion in a
model ecosystem yielded a calculated half-life of 15 to 16 days.
In a similar study, guthion yielded a half-life of one month at
pH levels less than 9 and a half-life of less than one week at
more alkaline pH's.
Interestingly, the degradation of organophosphates can be
inhibited by the presence of other synthetic organic chemicals.
Experiments were conducted with two detergent surfactants -
alkyl benzene sulfonate (ABS) and linear alkyl benzene sulfonate
(LAS). These experiments demonstrated increased persistence
for several organophosphate insecticides, especially parathion
and diazinon (304). As a result, highly polluted water may
exhibit accumulations or half-lives far beyond the normal for
organophosphates, which, as a rule, neither persist nor accumu-
late in the environment, but are removed entirely within a few
months.
Carbamate biocides are moderately soluble, ranging from 7
ppm for terbutol to 250 ppm for propham and averaging around 100
ppm (403). In general, they decompose easily and show little
197
-------
tendency toward adsorption on suspended material, but hydrolyze
readily. The hydrolysis is particularly pH dependent, virtually
ceasing below pH 5 (403) and increasing as the pH and tempera-
ture rise. High salt content affects the hydrolysis rate
inversely, slowing the rate as the salt concentration increases
(403). Carbamates photodecompose readily - increasingly so,
as the pH rises - and can be rapidly biodegraded under normal
circumstances (1170). Carbamates are not, then, persistent
in. normal water systems, lasting only a few days to a few weeks,
but remain as a stable compound in acidic waters (403).
Ionic biocides are a broad class embracing a variety of
chemical types and uses. They are all considered soluble in
water, with solubilities ranging from 100 to more than 1,000,000
ppm. Ionic biocides that are marginally soluble in pure water
have increased solubilities in natural waters high in humic
acid salts (403). With few exceptions, these biocides do not
accumulate or persist and, consequently, are seldom found in
high concentrations.
Ionic biocides are, however, strongly adsorbed onto soil
particles, all types of clay, humic matter, and organisms - in
short, onto anything with a partial charge or an ion exchange
capability (403). They are generally resistant to chemical
attack but photodegrade readily, except when adsorbed onto
particulate matter (403). Ionic biocides respond differently to
microbial attack, but are absorbed by many organisms. As a
result, they tend to concentrate in organisms and up the food
chain. Research on TCDD, an ionic herbicide residue, demon-
strated that accumulation was directly related to water concen-
trations (0.05 to 1,330 ppb) and averaged between 400,000 and
2,000,000 times the water concentration (627).
SYNTHETIC/ORGANIC CONTAMINANTS
Recently, there has been a great interest in identifying
synthetic/organic trace compounds in water supplies drawn from
rivers and in other water bodies receiving treated wastewater.
Although studies have been made of the concentrations found
various water systems, neither the environmental
the potential health effects to man of these
been studied to any great extent.
i n
pathways nor
substances have
Over 100 synthetic/organic compounds have been identified
in various drinking water sources. Thirty-six compounds were
found in the lower Tennessee River (Table 77 ), while 66 were
identified and Quantified in Mississippi River water at New
Orleans (Tabl» ':? '' "able 79 lists the results of organic
analyses of s?ver*i -it he" domestic water supply sources.
The many d i f t e >~ ~ n t
make generalizations di
ivpes of compounds under discussion here
f;,.jli regarding their environmental
198
-------
TABLE 77. ORGANIC COMPOUNDS IDENTIFIED
UP TO 1975 IN LOWER TENNESSEE RIVER (1323)
COMPOUNDS
Acenaphthene
Allylbenzoate
Anthracene
Benzene
Biphenyl
Butylbenzene
5-Chloro~2-Methylbenzofuran
p-Cresol
Diallyl Adipate
Dibutyl Phthalate
Diphenylacetylene
1,1-Diphenylethene
2,6-Di-Tert-Butyl-4-Methylphenol
Ethylbenzene
Ethyl o~Phthalate
Ethylstyrene
Ethylene Dimethylacrylate
Fluoranthene
Fluorene
Hexachlorobenzene
Indene
o-Methoxybenzoic Acid
2-Methylanthracene
2-Methylbiphenyl
4-MethyIdiphenylacetylene
Methyl Indene (2 isomers)
1-Methylnaphthalene
Naphthalene
p-Nonylohenol
n-Octyl-o-Phthalate
Pyrene
Styrene
1,2-Tetradecanediol
Toluene
3,4,4-Trimethy1-2-Hexene
Xylene
199
-------
TABLE 78.
ORGANIC COMPOUND IDENTIFICATIONS
NEW ORLEANS AREA WATER SUPPLY STUDY
(856 )
1
2
3
4
5
6
7
8
9
10
Highest Measured Concentration
yg/1
Compound
Acetaldehyde
Acetone
Alkylbenzene-Cg isomer
Alkylbenzene-Cg isomer
Alkylbenzene-C2 isomer
Alkylbenzene-Cj isomer
Alkylbenzene-C3 isomer
Alky1benzene-C3 isomer
Atrazine *
(2-chloro-4-ethylamino-
6-isopropylamino-
s_-triazine)
Deethylatrazine
(2-chloro-4-amino-
6-isopropylamino-
s_-triazine)
Carrollton
Water Plant
D-VOA
D-VOA
0.05
0.33
0.11
0.01
0.04
0.02
5.0
0.51
Jefferson # 1
Water Plant
NE
NE
ND
ND
0.03
ND
0.05
ND
4.7
0.27
Jefferson # 2
Water Plant
NE
NE
ND
ND
ND
ND
0.02
ND
5.1
0.27
200
-------
TABLE 78. (continued)
Highest Measured Concentration
11
12
13
14
15
16
17
18
19
20
21
Compound
Benzyl butyl phthalate*
Bromodichloroethane
Bromoform *
Butanone
Carbon disulfide
Carbon tetrachloride
bis-2-Chloroethyl ether*
Chloroform *»a
bis-2-Chloroisopropyl
ether *
n-Decane *
Decane-branched isomer
Carrol.lton
Water Plant
0.64
D-VOA
0.57
D-VOA
D-VOA
D-VOA
0.07
113
0.18
0.04
0.03
Jefferson # 1
Water Plant
0.81
NE
ND
NE
NE
NE
0.16
NE
0.05
ND
ND
Jefferson # 2
Water Plant
0.73
NE
ND
NE
NE
NE
0.12
NE
0.03
ND
ND
201
-------
TABLE 78. (continued)
Highest"Measured Concentration
Compound
22
23
24
25
26
27
28
29
30
31
32
Di bromodi chl oroethane
isomer
Dibromochloromethane *
Dibutyl phthalate *
2,6-Di-t-butyl-£-
benzoquinone *
Dichlorobenzene isomer
T,2-Dichloroethane a
Dichloromethane
Dieldrin **
Diethyl phthalate *
Di(2-ethylhexyl) phthalate *
Dihexyl phthalate
Carroll ton
Water Plant
0.33
1.1
0.10
0.22
0.01
8
D-VOA
0.05
0.03
0.10
0.03
Jefferson # 1
Water Plant
ND
0.30
0.16
0.19
D-RE
NE
NE
0.07
0.03
0.31
ND
Jefferson s? 2
Water Plant
0.63
0.60
0.19
0.23
ND
NE
NE
0.05
0.01
0.06
ND
202
-------
TABLE 78. (continued)
Highest Measured Concentration
yg/1
33
34
35
36
37
38
39
40
41
42
43
Compound
Dihydrocarvone
Diisobutyl phthalate *
Dimethyl phthalate
Dioctyl adipate
Dipropyl phthalate *
n-Dodecane *
Endrin **
Ethanol
£-Ethyl toluene *
£-Ethyl toluene *
1, 2, 3, 4, 5, 7, 7-
Heptochloronorbornene *
Carroll ton
Water Plant
0.14
0.59
0.27
0.10
0.07
0.01
0.004
D-VOA
ND
0.02
0.06
Jefferson # 1
Water Plant
0.06
ND
0.13
ND
0.13
ND
NYE
NE
0.04
0.03
0.05
Jefferson ? 2
Water Plant
0.07
ND
0.18
ND
0.14
ND
NYE
NE
0.02
0.03
0.05
203
-------
TABLE 78. (continued)
Highest Measured Concentration
Compound
44
45
46
47
48
49
50
51
52
53
54
Heptachloronorbornene
isomer
Hexachloro-1 ,3-butadiene *
Hexachloroethane *
Isophorone *
Limonene *
Methanol
Methyl benzoate
3-Methylbutanal
2-Hethylpropenal
n-Nonane *
n-Pentadecane *
Carroll ton
Water Plant
0.06
0.16
4.4
1.5
0.03
D-VOA
ND
D-VOA
D-VOA
0.03
0.02
Jefferson # 1
Water Plant
0.04
0..27
0.19
2.2
ND
NE
D-RE
NE
NE
ND
ND
Jefferson # 2
Water Plant
0.04
0.21
0.16
2.9
ND
NE
ND
NE
NE
ND
ND
204
-------
TABLE 78. (continued)
Highest Measured Concentration
ng/1
Compound
55
56
57
58
59
60
61
62
63
64
65
66
Tetrachloroethane
isomer
Tetrachloroethylene
n-Tetradecane *
Toluene *
1 ,1 ,2-Trichloroethane *
1 ,1 ,2-Trichloroethylene
n-Tridecane *
Trimethyl-trioxo-
hexahydrotriazine
isorner
Triphenyl phosphate *
n-Undecane *
Undecane-branched isomer
Undecane-branched isomer
Carroll ton
Water Plant
0.11
D
0.02
0.08
0.35
D-VOA
0.01
0.07
0.12
0.02
0.04
0.06
Jefferson # 1
Water Plant
ND
0.5
ND
0.10
0.45
HE
ND
ND
ND
ND
ND
ND
Jefferson # 2
Water Plant
ND
0.41
ND
ND
0.41
NE
ND
ND
ND
ND
ND
ND
205
-------
TABLE 78. KEY TO SYMBOLS USED IN TABLE
KEY TO SYMBOLS USED IN TABLE
Symbols used in column headed Compound
* While all compounds listed in the table were identified by one or
more methods, those marked with this symbol gained added confirma-
tion by gas chromatography retention time match with an available
standard of the compound.
** Compounds marked with this symbol gained further confirmation by
gas chromatography retention time match with available standards on
each of three different columns, polar and non-polar.
a The quantitative values for these compounds were obtained on
Volatile Organics Analysis by comparison with standards of known
concentration at the Water Supply Research Laboratory. Compound 18
was detected but not quantified in Tetralin extracts of Carrollton
water at Southeast Environmental Laboratory, but not in Tetralin
extracts of Jefferson Mo. 1 or Jefferson No. 2. The latter labora-
tory did not detect compound 27.
Symbols used in columns headed Highest Concentration Measured.
D-VOA These compounds were detected by Volatile Organics
Analysis - Bellar Technique only. Quantitative values
have not yet been obtained. Ttvis -method was performed
only on the Carroll ton water at the Water Supply Research
Laboratory.
D-RE These compounds were detected only on XAD resin
extracts in the specific water for which this symbol
is used. Quantitative values were not obtained from
the resin extracts. The compound may have been detected
and quantified by another method in one or both of the
other waters examined.
D In the one instance v/here this symbol was used the
compound was detected by both the Water Supply Research
Laboratory and Southeast Environmental Research Laboratory
but not quantified by either laboratory.
NE This symbol means not examined. It is used
exclusively for some compounds reported by the Water
Supply Research Laboratory. This laboratory did not
obtain samples of water from Jefferson No. 1 or Jefferson
No. 2.
206
-------
TABLE 78. (continued)
ND This symbol means the compound was not detected in
that specific water by any of the methods employed.
NY£ Compound 39 was confirmed in Carroll ton water carbon
chloroform extracts shortly before preparation of this
report. Jefferson No. 1 and Jefferson No. 2 extracts
have not yet been re-examined specifically for compound 39.
207
-------
TABLE 79. MOLECULAR CONSTITUENTS IDENTIFIED
IN NATURAL WATER SAMPLES (1323)
Constituent
p-Cresol
Diethylene glycol
Ethylene glycol
Glycerine
Glyci ne
Manni tol
Methyl-a-D-glucopyranoside
Methyl-B-D-glucophranoside
Sucrose
Xylitol
Urea
Inosi tol
0-Methyl inosi tol
Linoleic Acid
Oleic Acid
Palmitic Acid
Stearic Acid
2,2' -Bipyridine
Sample*
3
5
5
1,2,3,4,5
1
5
4
4
1,5
5
1,2
1,2,3,4,5
1,2,3,4,5
1,5
1,5
1,5
1
4
Concentration
mg/l
7
1
20
1-20
2
2
30
3
2
1
4
0.5-1
0.3-10
1
1
0.4
0.5
4
*1 - Lake Marion, 2 - Fort Loudon Lake, 3 - Holston River,
4 - Mississippi River, 5 - Watts Bar Lake
208
-------
fate. For instance, acetone is infinitely soluble in water;
chloroform is soluble to about 8,200 mg/t; carbon tetrachloride
is soluble to about 800 mg/t; and n-decane is insoluble. The
specific gravity of toluene is less than that of water, while
the specific gravity of carbon disulfide is greater. Acetalde-
hyde is readily metabolized since it is a natural metabolic
intermediate, but branched alkyls are almost impervious to
microbial attack.
The differences in man-made synthetic/organic compounds
exceed the similarities, but in general, these compounds are
persistent and resist microbial degradation. Beyond that
generalization, research has been too limited to discuss
specific compounds in detail.
Polychlorinated biphenyls (PCB's) are the only class of
synthetic/organic contaminants that have been studied in detail.
They are virtually insoluble in water, which, combined with a
high specific gravity and volatility, serves to keep solution
PCB concentrations low. However, PCB's are strongly adsorbed
onto suspended particulate matter and transported through the
water system. Because of their heavy, insoluble character and
sediment affinity, they tend to accumulate in bottom sediments.
A comparison of selected water and sediment PCB concentrations
from across the United States is presented in Table 80 .
PCB's are fairly stable in fresh-water systems, resisting
hydrolysis and chemical degradation, and are not amenable to
photodegradation (998). Theoretically, they should readily
vaporize from solutions, but this is prevented by their tendency
to sink or strongly adsorb onto suspended matter. Only PCB's
that are associated with floatables or oil slicks appear to
vaporize to any great degree. The lower isomers (four or fewer
chlorine atoms) are somewhat responsive to biodegradation, but
the degradation products are frequently more toxic than the PCB
itself (998). The higher isomers resist microbial attack.
PCB's are thus quite persistent in water/sediment systems, and
lifetimes of years or even decades have been postulated (998).
The continued presence of PCB's makes it inevitable that
they will enter the food chain. As they tend to accumulate in
lipid tissues in higher plants and animals, it has been
estimated that PCB's will concentrate up the food chain to as
much as 10' times the water concentration (999).
BIOLOGICAL CONTAMINANTS
An important pathway for certain communicable disease
transmission to man is the consumption of contaminated water.
Direct disposal of wastewater is the principal contamination
route. Land disposal of wastewater and sludge is not an
important pathway, as pathogenic organisms have limited mobility
209
-------
TABLE 80. PCB CONCENTRATIONS IN
SELECTED WATER COURSES (275)
State
Alaska
Arkansas
Cal ifornia
Connecticut
Hawai i
Georgia
Mary! and
Mississippi
New Jersey
Oregon
Pennsylvania
South Carolina
Texas
Virginia
Washington
West Virginia
Concentration
Water
ug/£
ND
ND
0.1, 0.1
0.1-0.2
ND
0.1
ND
0.1
ND
0.2
--
0.1-3.0
0.1
ND
ND
Concentration
Sediment
yg/kg
ND
20-2,400
20-190
40
ND
10-1 ,300
10-1 ,200
50; 170
8-250
15; 140
10-50
30-200
7.9-290
5-80
ND
10
210
-------
in soil and seldom migrate far enough to contaminate water
supplies (588). However, in contrast to their restricted
mobility in soils, biological contaminants are readily dispersed
and transported by receiving waters. Consequently, there is a
high potential for direct public contact through drinking or
recreational use. Wastewater treatment has diminished this
threat by reducing the number of organisms in the wastewater.
This, combined with natural pathogen mortalities, has greatly
lessened the outbreak of water-borne diseases attributable to
public water supplies.
The environmental factors that influence the survival of
pathogens in fresh water are, in most cases, similar to those
that prevail in marine systems. These factors will be discussed
in greater detail in the marine water section of this chapter.
Briefly, however, the chief factors influencing survivabi1ity
are temperature, pH, sunlight, toxins, predators, and lack of
nutrition, which affect pathogens to different degrees. An
examination of Table 81 reveals that pathogens may survive for
long periods of time and travel great distances before destruc-
tion by environmental factors.
Pathogenic bacteria are best adapted to survival in the
human body or to conditions resembling those found in the body.
Consequently, natural water systems are a hostile environment.
However, cool water is generally more hospitable than warm water
because of the depressed metabolism of both the bacteria and
their predators. Predatory organisms, especially in slightly
polluted waters, are a major contributor to bacterial die-off.
For instance, Barua (74) noted that Vibrio cholerae survived
one to two weeks in clean water as opposed to one to two days in
water with a large bacterial population.
Pathogenic bacteria also suffer from a lack of proper
nutrition in clean waters; low nutrient levels prevent reproduc-
tion. Since die-off rates exceed growth rates, the overall
population will decline. Other factors affecting die-off are
ultraviolet radiation in sunlight, pH extremes, natural anti-
biotics, and chemical toxins.
In contrast to bacteria, viruses do not multiply in water
and, therefore, their number in a water body can never exceed
the number introduced into that body by waste disposal. Typi-
cally, viruses are much more resistant to external environmental
factors (chemical content, pH, temperature, time, etc.) and
survive longer than bacteria (257, 380). It was long suspected
that algae could inactivate viruses through some process because
of low virus concentrations in algae-rich waters. However, it
is now believed that the high pH and dissolved oxygen in the
vicinity of algal blooms are responsible for the inactivation.
211
-------
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212
-------
Virus inactivation in lake water is further enhanced by
the presence of proteolytic bacteria which degrade the
viral coat (564). Coxsackie is particularly susceptible to
proteolytic bacteria, while polio virus is generally resistant
except to Pseudomonas aeruginosa (564). Otherwise, the mechan-
isms of virus removal are obscure. Table 82 reports survival
times for various enteric viruses in fresh-water bodies.
Transport mechanisms for pathogens include physical current
motion, organism motility, adsorption, ingestion, and aerosoli-
zation. As most pathogens readily adsorb onto suspended matter,
sediment pathogen concentrations may greatly exceed water con-
centrations. Filter feeding organisms, such as fresh-water
shellfish, tend to concentrate pathogenic organisms. Conse-
quently, shellfish can be a major factor in the spreading of
certain communicable diseases.
213
-------
TABLE 82. SURVIVAL OF ENTERIC VIRUSES
IN WATER (11)
Temperature
Type of Water Virus
River water Coxsackie B-3
ECHO 5
Polio 1
Coxsackie B-3
ECHO 12
ECHO 7
Coxsackie A-9
Polio 2
Polio 3
ECHO 5
Coxsackie A-9
ECHO 12
Polio 1
Polio 1
ECHO 7
Polio 1
Polio 3
Polio 3
Coxsackie A-2
Coxsackie A-2
Coxsackie B-5
Coxsackie B
Impounded fresh Coxsackie B-3
water Polio 1
ECHO 7
ECHO 6
Coxsackie A-9
Polio 1
Coxsackie ?
ECHO 12
Polio ?
ECHO ?
ECHO ?
Polio ?
Polio ?
ECHO ?
4-6
75/3
7/0
7/1
7/1
19/3
15/3
10/3
75/3
30/3
60/3
20/3
33/3
19/3
60/3
26/3
27/3
50/3
67/3
-
-
-
18/2
7/1
7/1
22/3
5/3
6/3
27/3
18/3
14/3
21/3
23/3
21/3
52/3
52/3
42/3
15-16
(Range of
8/3
.5
-
.7
-
-
-
15/3
8/3
15/3
-
-
-
45/3
-
-
18/3
7/1
-
-
24/1
-
.7
.5
-
-
-
-
-
-
-
-
-
_
-
-
°C
20-25
Days)
2/3
3/3
3/3
3/3
5/3
7/3
8/3
8/3
8/3
8/3
8/3
12/3
13/3
16/3
16/3
20/3
-
.3 7/2.1
5/2
47/2
-
-
3/3
3/3
4/3
5/3
6/3
6/3
8/3
6/3
10/3
12/3
20/3
21/3
22/3
24/3
214
-------
SALINE WATER
INTRODUCTION
The ocean has traditionally been the depository for a
wide variety of civilization's wastes. Previously, the ocean
was considered an infinite sink that was not affected by
the waters discharged into it. Only recently has man
recognized that wastewater effluents and sludges can have
an adverse effect upon ocean waters.
Wastewater (treated and untreated) and wastewater sludges
are discharged directly to the ocean offshore through sub-
marine outfall pipelines. Beach outfalls are occasionally
used for wastewater discharge, although this usually results
in contamination of the beach. Sludges are also barged and
dumped into the open ocean.
The potential significance of a given discharge varies
with the characteristics of the discharge and of the receiving
waters. For example, discharges into a bay or estuary are
much less diluted than discharges into the open ocean. The
location of the outfall (e.g., on a shelf or near a submarine
canyon) strongly affects contaminant mobility. Seafood
obtained from open ocean areas is different from that harvested
from bays or estuaries. Shellfish, which are usually taken
from bays or estuaries, tend to concentrate contaminants to a
greater degree than does open ocean seafood. Furthermore,
the physico-chemical relationships affecting solubility and
mass transport of contaminants differ considerably for
estuaries and open ocean receiving waters.
Consequently, the fate of wastewater contaminants in
marine environments is highly site specific. For any given
site, an understanding of the potential pathways from the
effluent back to man requires knowledge of the dilution
achieved, discharge location hydraulics, chemical and physical
composition of both the effluent and receiving waters, com-
position and hydraulics of other contamination sources,
physical and chemical properties of the sediments and their
interactions with the water column, and an understanding of
the food web relationships. Food web relationships are per-
haps the most important element in terms of ultimate pathways
to man, and have received the most publicity in recent years.
Public health-impairing contaminants behave in somewhat
the same fashion in marine water as in fresh water. Therefore,
215
-------
concepts discussed extensively in the preceding fresh-water
section of this chapter will be mentioned only briefly here.
In general, wastewater contaminants in the ocean can return
to man via two pathways: ingestion of contaminated fish and
shellfish, or body contact and incidental ingestion associated
with marine aquatic sports.
Research concerning the disposal of sewage effluents and
sludges to marine water systems is abundant, as can be seen
in Table 83 . The volume of literature is somewhat deceptive,
however, in that most of these studies concentrate on one
aspect of a pollution problem in a specific area; comprehen-
sive generalizations are usually not possible. The most
nearly complete picture of the pathways and fates associated
with wastewater discharges comes from a study of the Southern
California coastal area by the Southern California Coastal
Water Research Project (SCCWRP), which began in 1969.
Figure 21 shows the study area, which encompasses approximately
400 m of coast and 20,000 km2 of water surface area. This
coastal area receives municipal wastewater, sludge, surface
runoff, and other discharges from a population of approxi-
mately 11 million people. The effluent volume from municipal
treatment systems alone is approximately one billion gpd, of
which 20 percent receives secondary treatment while 80 percent
receives only primary treatment (1544).
Because of the completeness of the research, much of the
following discussion relies heavily on the SCCWRP data.
However, this data is specific to the Southern California area,
which differs from other ocean areas; hence, the data must be
applied with caution.
WATER QUALITY PARAMETERS
Since saline water is not used for drinking water, there
are no potential direct health problems from marine water
quality parameters. However, as in fresh surface waters,
there are some indirect consequences from the introduction of
suspended solids and organic carbon into the marine environ-
ment. Trace metals, bacteria, viruses, and other hazardous
contaminants are often found in association with suspended
particulates. Researchers in Southern California have
determined that more than 90 percent of the possible pollutants
discharged through one sludge discharge pipeline are associated
with particulate matter (925).
Discharged organic matter can have a radical impact on
the local environment which, in turn, will affect the dis-
tribution, transport, and availability of other contaminants.
The oxidation of organic matter, particularly in bottom
sediments, severely depletes dissolved oxygen concentrations,
216
-------
TABLE 83 . LITERATURE REVIEWED
PERTAINING TO SALINE WATER
Contaminant Reference Number
Water Quality Parameters
Ammonia 61, 221, 264, 358, 829, 925, 1002, 1102,
1125, 1149, 1207, 1541
BOD 226, 338, 925, 979, 1002, 1102, 1207,
1268
COD 925, 1268
Chlorides 225, 1125
Fluorides 1451
Nitrates 61, 86, 221, 225, 264, 358, 829, 882,
883, 1002, 1125
Nitrites 61, 86, 221, 264, 358, 829, 1002, 1125
Oil and grease 61, 925, 1102, 1207, 1231
Phosphates 61, 86, 221, 264, 358, 829, 882, 883,
925, 1101 , 1102, 1207
Suspended 459, 499, 925, 1125, 1171, 1172, 1541
solids
Total dissolved 1125
sol ids
Total organic 225, 559, 720, 979, 1125, 1268
carbon
Other (general) 42, 226, 551, 552, 700, 828, 925, 1067,
1102, 1125, 1207, 1307, 1310, 1544
Elemental Contaminants
Aluminum 28, 436, 636, 720, 968, 1125, 1128, 1160,
1161, 1266, 1380, 1413, 1459
Antimony 303, 436, 636, 720, 1125, 1128, 1160,
1161, 1266, 1380, 1525
Arsenic 303, 636, 637, 720, 1128, 1160, 1161,
1207. 1266. 1339, 1380, 1525
217
-------
TABLE 83.(continued)
Contaminant Reference Number
Barium 28, 636, 968, 1128, 1160, 1161, 1266,
1339, 1380, 1459
Beryllium 636, 1128, 1161, 1266, 1339, 1380, 1459
Boron 28, 636, 1113, 1128, 1160, 1161, 1266,
1339, 1380, 1459, 1525
Cadmium 53, 75, 115, 157, 227, 303, 365, 441,
524, 558, 592, 636, 713, 720, 756, 828,
923, 925, 967, 968, 978, 1102, 1117,
1121, 1125, 1128, 1160, 1161, 1207, 1208,
1378, 1433, 1459, 1502, 1541, 1544, 1562
Chromium 53, 75, 157, 212, 227, 232, 365, 611,
632, 636, 700, 713, 720, 828, 923, 967,
969, 1102, 1121, 1125, 1128, 1137, 1160,
1207, 1208, 1266, 1306, 1307, 1380, 1433,
1459, 1528, 1541
Cobalt 75, 115, 157, 303, 365, 407, 445, 636,
713, 782, 828, 894, 923, 967, 968, 1044,
1102, 1117, 1121 , 1125, 1128, 1160, 1161 ,
1266, 1339, 1380, 1459, 1538, 1544
Copper 75, 115, 157, 227, 232, 305, 365, 407,
445, 611, 636, 713, 720, 782, 828, 844,
894, 923, 967, 968, 1044, 1066, 1102,
1121, 1125, 1128, 1137, 1159, 1160, 1161,
1171, 1207, 1208, 1259, 1266, 1339, 1380,
1433, 1459, 1538, 1541, 1543, 1544, 1562
Germanium 636, 1128, 1160, 1161, 1266, 1339, 1380
Iron 28, 75, 115, 227, 232, 365, 611, 636,
720, 782, 828, 844, 968, 1113, 1117,
1121, 1125, 1128, 1160, 1266, 1339, 1371,
1380, 1459, 1502, 1544, 1562
Lead 75, 115, 157, 227, 365, 407, 441, 462,
524, 558, 611, 636, 713, 720, 828, 923,
967, 968, 1060, 1066, 1102, 1117, 1121,
1128, 1161, 1171, 1207, 1208, 1266, 1307,
1339, 1363, 1380, 1433, 1459, 1544
218
-------
TABLE 83. (continued)
Contaminant Reference Number
Manganese 75, 227, 232, 365, 445, 611, 636, 713,
720, 828, 844, 923, 967, 1102, 1117, 1121,
1125, 1171 , 1433, 1459, 1502, 1538
Mercury 12, 37, 227, 303, 305, 365, 368, 452,
462, 524, 610, 620, 636, 637, 645, 654,
760, 780, 795, 809, 967, 968, 987, 988,
1102, 1125, 1128, 1145, 1160, 1161, 1207,
1208, 1266, 1295, 1339, 1380, 1413, 1433,
1502, 1503, 1526, 1527, 1543, 1544, 1562
Molybdenum 303, 636, 720, 828, 925, 1128, 1160, 1161,
1266, 1339, 1380, 1459
Nickel 75, 115, 232, 365, 445, 636, 713, 720,
782, 828, 923, 967, 968, 1102, 1117, 1121,
1125, 1128, 1160, 1161, 1171, 1207, 1208,
1266, 1339, 1380, 1459, 1541
Selenium 636, 1114, 1128, 1160, 1207, 1266, 1339,
1380, 1525
Thorium 636, 1128, 1160, 1161, 1266, 1339, 1380
Tin 636, 828, 923, 1128, 1160, 1161, 1266,
1339, 1380, 1459
Uranium 636, 1128, 1160, 1)61, 1266, 1339, 1380
Zinc 75, 115, 157, 175, 227, 232, 303, 305,
365, 407, 441, 445, 524, 592, 611, 636,
713, 720, 782, 828, 844, 923, 968, 1066,
1102, 1114, 1117, 1125, 1128, 1160, 1161,
1171, 1207, 1266, 1307, 1339, 1380, 1433,
1459, 1538, 1562
Other (general) 37, 42, 75, 157, 228, 636, 720, 893, 923,
967, 1161, 1207, 1307, 1339, 1380, 1454,
1460
Biocidal Contaminants
Aldrin 403, 1237, 1418
Arsenated 1237
hydrocarbons
219
-------
TABLE 83.(continued)
Contaminant Reference Number
Chlorinated 304, 324, 365, 605, 804, 827, 874, 893,
hydrocarbons 977, 998, 999, 1022, 1023, 1059, 1065,
1066, 1068, 1102, 1125, 1132, 1146, 1162,
1170, 1207, 1231, 1237, 1307, 1412, 1532,
1535, 1536, 1539, 1543
ODD 403, 1022, 1125
DDE 403, 977, 1125
DDT 17, 365, 403, 420, 605, 872, 874, 894,
977, 1022, 1023, 1059, 1125, 1170, 1207,
1307, 1418, 1532, 1535, 1536, 1539, 1542,
1543
Dieldrin 365, 605, 1125, 1155, 1237, 1418, 1536
Endrin 403, 1418
Herbicides 806, 1125, 1412
Organophosphorus 1196
pesticides
Soil sterilants 73
Other (general) 42, 130, 304, 478, 806, 827, 1049, 1102,
1125, 1407, 1454
Synthetic/Organic 42, 58, 149, 235, 300, 304, 318, 319,
Contaminants 365, 402, 653, 768, 770, 807, 808, 879,
893, 903, 1025, 1125, 1126, 1132, 1454,
1533, 1539, 1558, 1564
Biological Contaminants
Adeno virus 350
Bacteria 18, 44, 45, 74, 365, 569, 656, 674, 858,
929, 938, 987, 1047, 1200, 1202, 1307,
1345
Clostridium 858
we!chi
220
-------
TABLE 83.(continued)
Contaminant Reference Number
Coliforms 17, 18, 44, 61, 197, 253, 254, 304, 310,
311, 316, 338, 612, 613, 656, 674, 675,
677, 695, 705, 938, 973, 989, 1019, 1020,
1021, 1047, 1087, 1200, 1202, 1222, 1223,
1224, 1226, 1231, 1268, 1275, 1307, 1413
Coxsackie virus 11, 252, 564, 905
(A & B)
ECHO virus 252, 357, 905, 1515
Escherichia 929, 1275, 1306, 1307, 1443, 1515
col i
Fecal 197, 399, 612, 1202, 1275, 1307
streptococci
Listeria 989
monocytogens
Mycobacteriurn 1306
Parasitic worms 700
Polio virus 11, 252, 564, 905
Salmonella 1234, 1306, 1307
Shigella 1306
Staphylococcus 350, 1306, 1307
aureus
Vibrio cholerae 255, 1020, 1306, 1316
Virus 11, 251, 252, 255, 311, 365, 405, 465,
468, 695, 877, 905, 926, 938, 1273,
1306, 1307, 1316, 1317, 1515
Other (general) 250, 311, 433, 858, 1125
221
-------
35°N
34aN
33°N
32°N
DRAINAGE
DIVIDE
^
120eW
119°W
118°W
117"W
Figure 21,
The Southern California Bight. Outfall
systems are (1) Oxnard City, (2) Hyperion,
Los Angeles City, (3) Whites Point, Los
Angeles County, (4) Orange County, and
(5) San Diego City. (1544)
222
-------
leading to anaerobiosis and a reducing environment. In such
an environment most dissolved metals will precipitate as
sulfides. The presence of dissolved organics enriches the
local environment in nutrients and leads to increased
planktonic growth and overall biological activity. Fish and
crustacean populations may also increase. Since these regions
of high organic content may also be high in hazardous elemen-
tals and organics, the chances of these fish and crustaceans
becoming contaminated are higher than for normal ocean water
populations.
Another problem associated with water quality parameters
is floatable material. Floatables from wastewater generally
can be classified as oil and grease, and methylene blue active
substance (MBAS). Again, the hazard is not with these
materials but with associated contaminants. Bacteria have been
shown to concentrate in floatables (365). Although research
has been scanty, it is probable that viruses, heavy metals,
biocides, and synthetic organics will also concentrate in
floatable materials. These floatables can be carried along
the water surface to the shore by wind, current, and wave
activity. Once deposited along the beach, these contaminants
become readily accessible to both humans and insect/mammal
vectors. Wind and wave action can also agitate floatables
to the point where aerosol-type particulates are released.
These airborne particulates can likewise be high in hazardous
contami nants.
As in fresh water, the addition of nitrogen and phosphorus
species to marine water can result in biostimulation.
Natural seawater concentrations of phosphorus and nitrogen
range from 1 to 100 ppb and 10 to 700 ppb, respectively.
Discharges of nitrogen and phosphorus to open ocean waters
are seldom a problem, as the ocean can assimilate these
nutrients. Discharges into bays, however, can lead to
accumulations of nitrogen and phosphorus with subsequent
biostimulation. Copeland's study of St. Joseph Bay, Florida
(264), found phosphorus concentrations as high as 10 times the
normal and gross biostimulation as evidenced by massive algal
growths. These growths decreased the aesthetic appeal of the
affected water and drove away many food and sport fish.
Dissolved oxygen levels also decreased, and the bay exhibited
evidence of eutrophication.
ELEMENTAL CONTAMINANTS
Marine elemental chemistry is similar in many respects
to fresh-water elemental chemistry. The same physical and
chemical principles apply, but the marine system is somewhat
more complex. Descriptions of the principles of solution,
precipitation, adsorption, etc., given in the fresh-water
223
-------
section of this report will not be repeated. Whenever the
complexities of the marine system Introduce modifications 1n
principles already discussed, these will be mentioned. In
general, however, this discussion will be limited to a simple
description of some of the Important components of the marine
system and their Interactions, with little mention of the
underlying principles.
The concentrations of elemental contaminants entering the
marine system via sewage effluent or sludge vary widely depend-
ing on the source of the effluent or sludge (Table 84 ).
Recent research has shown that 50 to 90 percent of these
elemental contaminants are associated with the particulate
fraction of the discharge (558, 967). Analyses of sediments
around outfalls seem to indicate that this elemental/sewage
particulate association does not last long in seawater,
although the mechanism of release is not well understood.
Among the possible explanations are:
• Oxidation of element-containing organic
particulates with subsequent release of the
element
• Oxidation of particulate element sulfides
• Surface desorption caused by high dilution ratios
_2
t Complexation with inorganic ligands such as C£
• Complexation with organic ligands possibly
resulting from the oxidation, of organic particulates.
The behavior of the elemental contaminants after their
release from the sewage particulates is scarcely better under-
stood. Table 85 presents the average natural marine concen-
trations and principal dissolved species for several of these
elements. It should be noted that these principal species
were probably identified in open ocean waters; dissolved
elementals may take different forms in coastal waters.
Furthermore the chemistry of the sewage-rich ocean waters near
an outfall can be radically different from the chemistry of
the open ocean.
Sewage-borne elemental contaminants are concentrated and
generally associated with suspended matter. At sludge outfalls
reducing conditions that prevail may lead to the formation of
insoluble sulfide compounds. The organics in the sewage can
lead to a reducing or only moderately oxidizing environment
where many metals -- such as zinc, lead, cobalt, cadmium, and
copper -- form insoluble carbonate compounds. Chromium
224
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225
-------
TABLE 85 . CONCENTRATION OF SELECTED ELEMENTS
IN MARINE WATER (1128)
Element
Aluminum
Antimony
Arsenic
Barium
Beryl 1 ium
Boron
Cadmium
Chromium
Cobalt
Copper
Germanium
Iron
Lead
Manganese
Mercury
Molybdenum
Nickel
Sel eni urn
Thorium
Tin
Uraniurp
Zinc
Concentrati on
(VQ/t)
1
0.3
2.6
20
6 x 10"4
4.5 x 103
0.1
0.5
0.4
3
0.07
3
0.03
2
0.2
10
7
0.09
<5 x 10"4
0.8
3
10
Principal Dissolved
Species
Al(OH)^3~y) • (H90)n
y 2 ' n
HAsOl , H9AsO«
+ 9
8a z
B(OH)3, B(OH)4
Cd+2, cdOH+, CdCl+
0 +^
CrO^, Cr °
Co + 2
Cu+2, Cu(OH)2
Ge(OH)4
+ +2
PbCl3> PbCl , Pb *
Mn + 2
HgCl42, HgCl2
Mo042
Ni+2, NiOH+, NiCl+
SeO^2
Th(OH)4
U02(C03)34
Zn + 2
226
-------
which is predominantly trivalent in sewage (632), forms in-
soluble hydroxides and oxides.
Morel et al. (967) and Hendricks (558) indicated that
less than 10 percent of the particulate elements settle
in the immediate outfall area. Rather they are carried away
by current action, and the dilution and oxidation potential
are increased significantly. When thus spread and diluted,
many of the precipitates formed near the outfalls will
dissolve, e.g., sulfides will oxidize to soluble sulfates,
and carbonate compounds will dissolve due to the action of
competing ligands.
Once in solution in open water, dissolved elements
are subject to a wide variety of competing mechanisms that
determine whether a given elemental remains in solution or
is removed. In highly dilute, oxygenated seawater, most
elements are in solution as:
• Cations: Ba+2, C0+2, Cu+2, Ni+2
• Oxo anions: HAs04"2 , Cr04"2, Se04"2, Mo04~2
• Chloride complexes: CdCl+, PbClg", HgCl4"2, NiCl+
• Hydroxyl complexes: B(OH)3, B(OH)4§ Cu(OH)2
• Carbonate complexes: L^KO.,).," .
Iron and manganese are major exceptions to the solubility
rules; both are more soluble in their lower valence states,
and as a result, are mobilized in reduced sediments and enter
solution as the divalent cations.
When divalent iron and manganese migrate to more highly
oxygenated waters, they oxidize to form insoluble hydrous
oxides that precipitate. These hydrous oxide particulates
have a high affinity for other ions and act somewhat as
scavengers. They are particularly effective at adsorbing
cobalt, nickel, cadmium, zinc, silver, selenium, and lead
but will also remove some chromium, copper, uranium, and
molybdenum.
If these hydrous iron and manganese oxides settle into
reducing sediments, the iron and manganese are again released
into solution. The accompanying ions, however, are usually
locked into the sediments as the extremely low solubility
carbonate or sulfide compound.
If these hydrous oxides settle into oxidizing sediments,
the adsorbed ions generally become incorporated into the
227
-------
precipitate. The iron and manganese oxides accrete to form
the ferro-manganese nodules that cover much of the ocean floor.
These nodules are invariably enriched in lead, cobalt, nickel,
zinc, cadmium, etc. A major exception to this behavior is
chromium, which is often carried with the hydrous oxides as
the Cr(III) oxide. The redox conditions around the iron/
manganese minerals are such, however, that the Cr(III) is
oxidized to the highly soluble Cr(VI), which is then released
back into solution. Most of the elementals, once they become
incorporated into the sediments, are effectively locked
away from further reaction with the seawater, barring bio-
logical activity or a major change in redox potential (Table
86 )•
Biological activity also plays an important role in the
transport of elemental contaminants. Biologically related
mechanisms take several forms: uptake and biomagnification
by organisms, biological transformation, sedimentation
associated with dead organisms or fecal matter, and complexes
with biologically related organic ligands.
Generally, elemental contaminants become a sociated with
marine organisms via one of several pathways. Some trace
metals (e.g., cobalt, iron, copper, zinc) can be incorporated
directly into metabolic pathways. Other trace metals
(notably mercury) exhibit 1ipid-preferential solubilities
whereby they concentrate in the fatty cells of small organisms,
the gill tissues of fish, or the digestive organs of filter
feeders. Metals can adsorb onto the surface of plankton
or can be transferred to higher marine organisms (and ul-
timately to terrestrial organisms and man) via ingestion of
contaminated food organisms. Two other pathways, bioaccumula-
tion and biomagnification, are discussed in more detail in
the "Pathways to Man" section of this report, under "Fish"
and "Shellfish."
Incorporated elemental contaminants can be returned to
the sediments in fecal material or in dead organisms. If
this occurs in a biologically active area, the elementals
can be ingested or sorbed by other organisms or simply re-
leased again to solution; otherwise they will probably be
incorporated into the sediments. Sediment elementals (whether
there by chemical or biological means) are subject to
ingestion by bottom feeders or transformation by sediment
bacteria. Biotransformation is probably an important trans-
port mechanism. For example, inorganic mercury and its
compounds, chemically immobile in sediments, are readily
absorbed by many bacteria. These bacteria, possibly via a
detoxification mechanism, convert the inorganic mercury to a
organomercurial. The organomercurial, usually methyl mercury,
is far more mobile than its inorganic counterpart. It is
228
-------
TABLE 86 . NATURAL TRACE METAL CONCENTRATIONS (MG/DRY KG)
AS REPORTED FOR SURFACE SEDIMENTS FROM SEVERAL
PARTS OF THE WORLD OCEAN (365)
Trace
Metal
Silver
Cadmium
Cobalt
Chromium
Copper
Iron
Mercury
Manganese
Nickel
Lead
Zinc
So.
Calif.8
1.0
0.4
7
46
16
2.5%
0.06
320
14
8
63
Other
Nearshore
2f
59
309
7h, 481
1-5%J>
0.04k
580f
15h, 551'
15h, 201
951'
Organic
Rich Cont. Pacific
Shelfb Basins0
4-195
4-93
18-129 82-686
0.4-5.5%
0.09-0.86
42-1,075
35-455
3-32
18-337
Manga-
nese d
Nodules
3,000
5,000
14%
19%
4,000
1,000
400
Outfall
Max.8
21
79
11
1,000
670
3.7%
4
380
45
490
2,400
Averages of natural values at the five major sewage outfalls in the
Bight. However, mercury value is estimated natural subsurface con-
centration in Santa Barbara Basin.
Organic-rich diatomaceous muds, S.W. Africa shelf.
c Sample No. 21-29, Pacific Ocean Basins.
Average composition of ferromanganese minerals from the Pacific Ocean.
6 Maximum surface sediment concentration measured off Palos Verdes
Peninsula sewage outfalls.
Caribbean sediment, Station 50.
9 Japanese Island sediments,
Gulf of Paria, Caribbean sediments.
Atlantic nearshore sediments.
Gulf of Paria, Caribbean sediments.
Washington Shelf, N.E. Pacific.
229
-------
easily released into the water column and is even more lipid
preferential than inorganic mercury. Only mercury sulfide
compounds and mercury bound to organic sulfhydryl groups are
relatively unaffected by bacteria. This type of behavior has
also been postulated, or demonstrated in the laboratory, for
arsenic, selenium, antimony, lead, cadmium, and other
elementals.
Organic ligands from decaying organisms or excretory pro-
ducts also may play a role in the solubility of many elements.
Examples are the copper-heme, iron-heme, and cobal t- vi tami n B-,~
complexes. All are natural complexes and are extremely
stable in marine environments. Several metals, such as mercury
and nickel, show a high affinity for organic ligands and
particulates. These element-organic ligand complexes will
stay in solution until the organic nqand decays, the complex
is ingested or sorbed, or the elements can form a more stable
compound.
The ocean as a whole is in a relatively steady state or
equilibrium condition, i.e., the rate of input of natural
materials roughly equals the rate of sedimentation. Conse-
quently, the concentrations of most elemental contaminants
in seawater are constant. Generally speaking, those elements
present in highest concentrations are the least reactive and
least likely to sediment. The converse is equally true.
Moreover, the elements considered hazardous to man are among
those present at extremely low concentrations. Therefore,
it is difficult to see how man's additions can have a serious
impact on the levels of these contaminants in seawater. Of
course, man has been adding these metals to seawater for
decades; but, as was just mentioned, these metals are among
the most reactive and most likely to precipitate. The
cations, an ions, and ligands most responsible for this pre-
cipitation are present in excess and are much more readily
replenished than are the more hazardous elements. Conse-
quently man's additions can be removed relatively rapidly
without seriously affecting marine equilibrium.
While this may sound comforting, it does not take into
account the fact that sludge and wastewater disposal occurs
in coastal waters where exchange with open ocean waters may
be limited. Ocean waters form an open, steady-state system;
coastal waters form a partially closed system that is subject
to varying inputs from the land. Thus coastal waters are
the most biologically active of the ocean's areas; conse-
quently, they a^e far more readily upset by wastewater addi-
tions than ;s t'*e o-.,ean as a whole.
It would u. i- e < ~: r ?", e 1 y useful to be able to describe the
relative impact d'ffcrent wastewater element concentrations
230
-------
would have on coastal equilibrium models. Unfortunately,
as discussed earlier in this section, there is a distinct
lack of uniformity among various coastal regions that makes
the task virtually impossible. Even the SCCWRP (the most
complete characterization of a specific region) does not
adequately address the problem of the diverse effects of
different wastewater levels. Suffice it to say that, in
general, wastewater disposal in coastal waters will increase
the concentrations of elements in the water, sediments, and
marine organisms.
BIOCIDAL CONTAMINANTS
The behavior of the biocidal contaminants in marine
water systems is largely the same as in fresh-water systems.
The mechanisms of microbial and chemical degradation, photo-
degradation, sediment and particulate adsorption, volatiliza-
tion, and biological uptake are physically the same. The most
important difference is the reduced solubility of many of the
biocides in the more highly ionic seawater. This will lead
either to increased sedimentation or increased concentration
in surface films, depending on the density of the particular
b i o c i d e .
Biocides in marine systems have not been studied as
extensively as biocides in fresh-water systems. Most of
the research on the former is concerned with uptake by or
effects on aquatic life. Little has been done to characterize
the physico-chemical transport and removal mechanisms active
in seawater. Furthermore, what research has been conducted
has been limited largely to DDT.
Table 87 lists the reported concentrations of DDT,
PCB, and dieldrin in selected effluents and sludges dis-
charged to the ocean off Southern California. These effluents
contain both domestic and industrial sewage and, in some
cases, are not treated beyond advanced primary treatment.
The chlorinated hydrocarbons are not metabolized to
any extent in seawater, polluted estuaries, or bottom sedi-
ments (1059). Rather the bottom sediments, particularly those
near the outfalls, act as a sink where these biocides may
persist for years and even decades (874).
Suspended or dissolved chlorinated hydrocarbons are
subject to uptake by algae. A number of researchers have
determined that algae are extremely efficient accumulators
of these biocides and thei*- residues (804, 1155, 1418). The
algae do not degrade the biocides but store them. The
biocides can be released if the algae are placed in clean
seawater. Otherwise, the algae retain the biocides until the
231
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algae are either eaten or die. in which case the biocides
will sediment with the decaying organic matter.
Chlorinated hydrocarbons in surface films are subject
to photodegradation, but even this may be more of a problem
than a solution. For instance, Rogers and Landreth (1170)
demonstrated that DDT could photodegrade to PCB's.
Although the sediments act as a sink for many biocides,
they do not permanently remove them from the environment.
On the contrary, contaminated sediments will act as a source
for the biocides long after major reductions have been made
in the dominant inputs. Bottom-feeding fish placed in an
otherwise clean environment (clean seawater with no biocide
inputs but with contaminated sediments) soon show significant
increases in tissue biocide content (1539). This problem
is limited largely to PCB's and the persistent chlorinated
hydrocarbons. Carbamates, organophosphates, and ionic
biocides all hydrolize rapidly in seawater, with half-lives
of hours or days, seldom exceeding a few weeks.
SYNTHETIC/ORGANIC CONTAMINANTS
Most of the research conducted to date on organic con-
taminants in the marine environment has been concerned with
PCB's and the petroleum hydrocarbons associated with oil
spills. Little has been done to determine the environmental
fate of the plethora of other synthetic organic compounds
released to the ocean. Because of the nonpotability of
seawater, little concern has been expressed regarding the
fate of the synthetic/organic compounds therein. The lipid
solubility of some of these compounds, however, suggests
that some research is in order, though the sheer number of
compounds may discourage meaningful research.
Polychlorinated biphenyls are one group of synthetic/
organic contaminants that has been studied in depth. PCB's
are highly insoluble in seawater, adsorb rapidly and strongly
onto suspended matter, and concentrate in surface oil slicks
or sediment. Adsorbed PCB's will ultimately diffuse into the
sediments, while those PCB's concentrated on the surface
photodegrade to lower isomers that hydrolize relatively
easily (1132).
The research on petroleum hydrocarbons is only partially
useful. Municipal wastewater ranks low among marine petroleum
hydrocarbon sources; natural seeps, oil spills, aerial fall-
out, industrial discharge, and deliberate dumping all con-
tribute more petroleum hydrocarbons to the sea. However, as
the behavior of many organic compounds in seawater is similar,
theoretical pathways can be suggested.
233
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In general, any organic compound that can occur naturally,
can be metabolized and effectively removed from solution.
Totally synthetic compounds may or may not be metabolized,
but those structurally similar to natural compounds are more
amenable to microbial attack than those with no natural analogs.
Increasing substitution, particularly with halogens, increases
resistance to biodegradation.
Most organic compounds are degradable via photolytic
oxidation or chemical attack. In some cases, e.g., plastics
or chlorinated hydrocarbons, half-lives of years are not
unusual. Most organic compounds, however, degrade more
quickly; half-lives of hours to perhaps a few weeks are more
common.
Chemical degradation may take the form of hydrolysis or
reaction with another chemical species in the water. Hydroly-
sis is a reaction with water and usually consists of adding
a hydroxyl group or breaking an ester-type linkage. The
number of chemical reactions possible are almost limitless.
Furthermore, there is no guarantee that reaction or hydroly-
sis products will be any less harmful or more biodegradable
than the parent compound. For instance, azo dyes are readily
attacked by hydrogen sulfide to give carcinogenic, non-
biodegradable benzidines (1126).
Photodegradation is limited to surface films, as
ultraviolet radiation will not penetrate beyond a few
centimeters. Consequently those organic compounds dissolved
in surface films, adsorbed onto floatables, or less dense than
water are most amenable to photolysis.
Again, in general, municipal effluents are not responsible
for most organic contaminants in the oceans. Furthermore,
most organics discharged in such effluents, if not degraded,
remain in the general vicinity of the outfall. Also, the
wide variety and small quantities of most organic contaminants
make it unlikely that any one will become a hazard in recrea-
tional waters or food organisms. Unfortunately this does not
take into account possible synergistic effects of contaminant
combinations nor the postulated zero threshhold value for
some carcinogens.
BIOLOGICAL CONTAMINANTS
Both untreated and partially treated wastewater and
sludge are carriers of a variety of pathogenic organisms.
Furthermore, since sewage destined for ocean disposal is
often given minimal treatment, pathogen concentrations in
effluents and sludges can be extremely high (44).
234
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There are five principal types of human pathogens, and
many of the disease-producing organisms therein can be trans-
mitted via water routes. The five types and some of the more
prevalent waterborne diseases are:
• Protozoa (amebiasis, balantidiasis)
• Nematodes (angiostrongyliasis )
• Platyhelminths (clonorchiasis)
t Viruses (viral hepatitis, viral meningitis,
viral respiratory diseases)
t Bacteria (coliform diarrhea, cholera, 1eptospirosis ,
salmonellosis, shigellosis, typhoid fever)
There has been little research on the behavior of
pathogenic protozoa, nematodes, and platyhelminths in marine
systems. They are not as prevalent in industrialized
societies nor generally as serious from a public health stand-
point. Furthermore, they are largely host specific with few
marine hosts and usually a problem only in fresh-water systems.
Bacteria and viruses are much better represented in the
literature. Both are virtually ubiquitous, and the diseases
they cause are more prevalent with more serious outbreaks.
A number of researchers have attempted to assess the fate of
these human pathogens in the marine environment. Unfortunately,
many of the studies have been done in the laboratory with
simulated seawater, and the results may not always be appli-
cable to real systems. At best, these laboratory studies can
only evaluate one or two factors, so the effects of factor
combinations and synergisms must remain largely a mystery.
Consequently, those factors primarily responsible for pathogen
die-off or transport in the sea have not been isolated
definitively. Table 88 presents reported marine survival
times for viruses and selected bacteria.
The greatest proportion of pathogenic bacteria in
municipal effluent is associated with particulate matter (44).
As a result, ordinary sedimentation accounts for the initial
removal of many bacteria. However, some bottom-feeding fish
and shellfish can accumulate bacteria (e.g., Clostridium
perfringens ) and provide favorable living condi ti bris ("8 58).
Consumption of raw or partially cooked seafood taken near
outfalls then becomes a problem.
Bacterial populations can increase near outfalls as
well. The relatively higher nutrient content of the water
near an outfall provides a much better growth media than does
normal seawater, with a resulting increase in bacterial
235
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TABLE 88 . BIOLOGICAL SURVIVAL' TIMES (DAYS)
IN MARINE ENVIRONMENTS
Vi ruses
Viruses
Viruses
Salmonel la
E. Coll
2-130
<28-130
15-88
>25
58-69
>25
> 4
75
Source
468
905
11
1307
1234
1307
929
1306
populations (1200, 1515). In general, though, normal seawater
Is considered bacterlostatlc, and d1e-offs rather than pop-
ulation Increases prevail in most of the ocean. These die-
offs are attributable to a variety of factors, including:
• Sensitivity of many bacteria to oceanic extremes
of pH and temperature (365, 674);
• Severe disruption of normal osmotic cell balance
due to marine salinities (365);
• Bacteriostatic effects of solar radiation on
any organisms near the surface (365);
• Bacterial predators flourishing near outfalls
(365, 929);
• Secretion by marine plankton of a fairly potent
antibiotic (44, 45); and
• Normal seawater nutrient levels that are too low
to support bacterial growth (345, 1200, 1515).
Consequently, bacterial die-off is generally rapid away
from the influence of the outfall, except for those bacteria
that may be ingested by fish or shellfish.
Bacterial populations will stay high in areas where
effluent discharges are continuous, particularly if there is
little current mixing. As a result, many bays and estuaries
can exhibit pathogen concentrations far beyond normal expecta-
tions, and recreational use of the water may become hazardous.
236
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Pathogenic bacteria tend to concentrate In surface films
and floatables. Aubert et al. (44) found concurrent coliform
concentrations of <1/100 ml in the water, 3 to 300/100 m£ in
surface films, and 3,500 to 21,000/particulate in floatables.
This raises the possibilities of aerosolization of bacteria
or surface transport of bacteria to recreational beach areas.
Viruses are far more resistant to seawater than are
bacteria. The relative time required for 90 percent inactiva-
tion of viruses in seawater has been estimated to be three to
six times that for coliform bacteria. Water temperature seems
to be the most important factor in virus inactivation; water
temperatures of 25°C and lower yield survival times in months
(252). Studies in warm waters, such as off Tel Aviv, have
shown 90 percent inactivation in less than 48 hr (365). Die-
off times are also largely virus specific, with several
studies showing that coxsackie survives best, followed by
ECHO virus and then polio virus (252, 905).
Although the viruses survive longer than do bacteria,
they do not have the tendency to increase in population as
do bacteria (1515). Viruses require living hosts to replicate
and are very host specific. Human viruses generally do not
reproduce in nonhuman hosts.
Viruses can, however, accumulate in marine organisms.
Fish and shellfish can ingest viruses and essentially store
them. Viruses survive quite well in a variety of shellfish
without harming their hosts (365, 905). Viruses can also com-
centrate on particulates and floatables with the same con-
sequences as for bacteria.
237
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LAND/GROUNDWATER
INTRODUCTION
Wastewater treatment sludges have traditionally been
disposed to land, and wastewater effluents are also increas-
ingly being disposed to land. Land disposal of sludges and
effluents, if improperly managed, can result in a potential
public health hazard. This hazard can be brought about
directly through surface contamination of food crops, or
indirectly through contamination of groundwater supplies,
consumption of livestock fed with contaminated crops, and
surface runoff affecting water supplies, aquatic life, or
bodies of water used for recreational activities. Aerosols
from spray application of effluents and/or direct vaporiza-
tion of volatile contaminants may also pose a direct or
indirect threat to human health. The major potential prob-
lem is the contamination of groundwater resources, which sup-
ply a significant percentage of domestic water consumption
in the United States.
Many variables affect the potential for a public health
hazard from land disposal. These include:
• The characteristics of the sludge or wastewater
di sposed
t The rate of waste application
• The hydrogeological characteristics of the disposal
site
• The method of disposal, e.g., crop irrigation, land
spreading, percolation ponds, sanitary landfill, etc.
• Proximity of public access
• Utilization of groundwater, crops, etc., potentially
affected
• Local climate.
The above listed variables are not mutually exclusive;
they interact in many ways to influence the rate and extent of
transport of contaminants from each source and to influence the
importance of each potential pathway. It is beyond the scope
of this report to delve deeply into the science of waste land
238
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application technology. This subject is being researched heavily
by the EPA and other agencies to establish guidelines for the
safe land disposal of effluents and sludges. This report will,
however, discuss the current knowledge about potential public
health problems.
Literature reviewed concerning wastewater and sludge
disposal to land (tabulated in Table 89) has centered on ground-
water elemental and biological contaminants, and water quality
parameters providing plant nutrients, such as nitrogen and
phosphorus forms. The increased use of wastewater application
to land has created a need for increased research into the
dynamics of wastewater contaminants in soil systems. It is
necessary to study the capabilities of these systems to destroy,
neutralize, remove, concentrate, or otherwise affect applied
wastewater contaminants.
A number of factors determine the degree to which ground-
water may be contaminated by wastewater or sludge that is
applied to land. Depth to the groundwater table and distance
to an extraction point affect residual levels of phosphorus,
bacteria, and other constituents for which removal appears to
be a function of travel distance. Soil characteristics, native
groundwater quality, assimilation capacity of the aquifer, and
method of waste application also determine groundwater degrada-
tion and consequent health problems (1214). Cation exchange
and adsorptive capacities important in the removal of metal ions
and viruses, and of trace organics and solids, respectively are
determined by soil composition. Porosity regulates infiltration
rates to some extent, affecting contaminant residence time in
surface layers. Residence time may, in turn, determine aerobic
or anaerobic conditions.
Total groundwater volume cannot necessarily be considered
an effective diluting agent. Uniform diffusion of recharged
water cannot be guaranteed, and water quality may vary consider-
ably both in area and in depth.
WATER QUALITY PARAMETERS
Research in this area has been principally concerned with
nitrogen and phosphorus forms entering groundwater as a result
of land application of wastewater and sludge. Removal of
suspended solids from wastewater effluent has also received
attenti on.
The problems and transformations associated with nitrogen
forms in soils are relatively well known. Organic nitrogen
and ammonia, when applied to soils under normal aerobic condi-
tions, are rapidly converted by nitrifying bacteria to nitrite
and nitrate. Both of these anionic forms move through soils in
239
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TABLE 89. LITERATURE REVIEWED PERTAINING
TO LAND/GROUNDWATER
Contaminant
Reference Number
Water Quality Parameters
Ammonia
BOD
COD
Chiorides
Fluorides
Ni trates
Ni tri tes
Phosphates
Suspended solids
Total dissolved
solids
Total organic
carbon
Other (general )
Elemental Contaminants
Aluminum
Arsenic
Boron
Cadmi urn
26, 47, 134, 218, 377, 479, 704, 705,
755, 1118, 1214, 1323, 1328, 1484
26, 134, 1140, 1215, 1233
26, 134, 1148, 1214, 1484
26, 47, 352, 1018, 1215, 1419, 1440,
1484
134, 1119, 1173
26, 47, 134, 218, 272, 377, 479, 690,
704, 705, 776, 798, 1118, 1119, 1167,
1214, 1233, 1323, 1328, 1419, 1440,
1484
47, 218, 479, 705, 798, 1118, 1119,
1284, 1432, 1484
47, 68, 134, 352, 595, 690, 755, 1214,
1299, 1419, 1484
47, 84, 134, 352, 415, 1140, 1214,
1215, 1233, 1388
26, 134, 352, 1215, 1304, 1323
134
106, 169, 393, 916
378, 634, 776, 1108, 1323
1323, 1419
134, 378, 736, 797, 1007, 1024,
1327, 1484
134, 401, 690, 776, 1323
240
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TABLE 89 (continued)
Contaminant
Reference Number
Chromium
Cobalt
Copper
Iron
Lead
Manganese
Mercury
Nickel
Selenium
Zinc
Other (general )
Biocidal Contaminants
Chlorinated
hydrocarbons
DDT
Dieldrin
Herbicides
Other (general)
Synthetic/Organic
Contaminants
Biological Contaminants
Bacteria
776, 1419
634, 776
134, 378, 401, 634, 690, 736, 776,
797, 1007, 1024, 1323, 1327, 1495
26, 47, 634, 736, 776, 782, 797, 939,
1007, 1024,.1214, 1285, 1323i 1327
39, 134, 634, 776, 782, 1323
378, 634, 736, 776, 782, 797, 1007,
1024, 1323, 1327
128, 134, 401
396, 401, 634, 690, 798, 1141, 1423
1323
134, 378, 634, 690, 736, 776, 797,
1007, 1024, 1323, 1327
634, 661, 690, 1173, 1284, 1440
401
401
401
1014, 1417, 1474
1173, 1323, 1326
934, 1384
428, 619, 860, 892, 1105, 1175, 1210,
1304, 1323, 1388, 1419
241
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TABLE 89. (continued)
Contaminant Reference Number
Conforms 47, 134, 155, 401, 776, 917, 1129,
1140, 1175, 1215, 1233, 1323, 1440,
1484
Coxsackie virus 322
(A&B)
Fecal 1210
streptococci
Parasitic worms 1323
Polio virus 322, 774
Protozoa 428, 1304, 1323
Salmonella 401, 1083, 1323, 1440
Virus 134, 322, 352, 428, 892, 1083, 1175,
1210, 1215, 1304, 1323, 1479, 1548
Other (general) 619, 776, 917
percolating water with little difficulty. Under anaerobic soil
conditions, on the other hand, the process of conversion to
nitrite and nitrate is inhibited. Ammonium ions and free
ammonia persist and are held near the soil surface by adsorption
onto soil particles, by cation exchange reactions, or by fixation
in clay lattices.
In an acidic environment, nitrite has been found to react
with secondary amines to produce nitrosamines. These compounds
have recently been labeled carcinogenic, teratogenic, and
mutagenic. The health hazards associated with nitrite and
other forms of nitrogen in drinking water and crops have been
delineated by the U.S. Department of Agriculture (1432).
The most definitive study of nitrogen removal by land
application of wastewater effluent was conducted by Bouwer, et al.
at Flushing Meadows, Arizona (134). They found that short flooding
periods (two days flooding followed by five days drying) did not
provide sufficient time to develop the anaerobic conditions for
nitrate denitrification. A longer flooding period of ten days
followed by two weeks of drying proved to be more favorable.
242
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With this schedule, oxygen in the soil was depleted during
flooding, causing nitrogen (in the ammonium form) to be adsorbed
by the clay and organic particles. Flooding was stopped before
the cation exchange complex in the soil was saturated with
ammonium. Upon drying, oxygen entered the soil, and ammonium was
nitrified under aerobic conditions to nitrate. Concurrently,
some of the nitrate formed was denitrified - in micro-anaerobic
pockets in the otherwise aerobic upper soil zone - to nitrogen
gas that escaped to the atmosphere. When flooding was resumed,
if the basins were immediately flooded to a depth above 1 ft,
the nitrates were quickly leached out of the top few feet of
the soil to groundwater. However, if initial flooding was
shallow (a few inches deep), the lower head allowed a low infil-
tration rate, a larger nitrate retention time in the micro-
biologically active soil zone, and further denitrification. At
these lower initial hydraulic loading rates, nitrogen removals
were as high as 80 percent. If high application rates were
consistently maintained, nitrogen removal was only 30 percent
with a peak nitrate surge to the groundwater after the start of
each new flooding cycle.
A study by Preul (1119) in 1966 provided the following
observations of the movement and conversion of nitrogen in soil
and the potential dangers of nitrate contamination:
1. Biological oxidation is the dominant mechanism
affecting ammonia nitrogen as it passes through
the soil. This action initially occurs at a high
rate and to a large extent within several feet of
the point of release of the septic tank effluent,
if soil conditions are well aerated.
2. Nitrate contamination of groundwaters is a serious
threat from shallow soil adsorption systems. High
concentrations of ammonia nitrogen in septic tank
effluents are quickly nitrified to high concentra-
tions of nitrate, which pollute the groundwater.
Dilution from groundwater or soil moisture and
possibly denitrification aid in the deterrence of
nitrate.
3. The effectiveness of adsorption in deterring the
travel of nitrogen is limited because of the
rapid conversion of ammonia to nitrate. Laboratory
experiments have shown that ammonium can be readily
removed in soil by adsorption but, under aerated
soil circumstances, nitrification of these ions
occurs before the flow can contact a sufficiently
effective volume of soil.
243
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Similarly, results of a study by Chapman et al. (218) have
shown that, in Texas, irrigation with a sewage effluent was a
potential source of nitrate pollution of the local groundwater.
The results indicated that nitrification of ammonia nitrogen
in the effluent is rapid and complete, taking place within the
top 3 ft of soil. It was concluded that substantial amounts
of nitrate would not be fixed by the soil and that, at a 3-in
per week application rate, appreciable amounts of high nitrate
water would percolate to the groundwater. Selective crop
production of grains and grasses having high nitrogen uptakes
(corn, bermuda grass, oats) was deemed the most effective method
of protecting the groundwater.
Short daily flooding schedu
efficient nitrification to nitra
reported data from test basins i
Rio Hondo spreading grounds near
basins, equipped to collect wate
received treated wastewaters on
period of flooding and a longer
ensured completely aerobic condi
almost all the nitrogen had been
at depths of 8 ft.
les evidently result in highly
te. McMichael and McKee (892)
n the Whittier Narrows and the
Los Angeles, California. These
r at 2-, 4-, 6-, and 8-ft depths,
a daily basis using a short
period of drying. This cycle
tions. Studies indicated that
converted to the nitrate form
Intermittent and continuous spreading of secondary effluent
at the aforementioned Hyperion Treatment Plant (1233) resulted
in the nitrogen transformations shown in Table 90 .
TABLE 90 . NITROGEN TRANSFORMATIONS RESULTING FROM
DIFFERENT SPREADING TECHNIQUES (1233)
I.
Under continuous
spreading:
Effluent applied
Organic-N Ammonia-N Nitrate-N
7.1-8.5 15.5-17.5 0.2- 0.8
II.
Percolate from 7-10 ft
below ground (mg/£)
Under intermittent
spreading:
Effluent applied
1.2-2.7
8.7-18.5 5.2-18.1
Organic-N Ammonia-N Nitrate-N
2.3-3.6 4.9-27.9 0.1-14.2
Percolate from 7-10 ft
below ground (mg/£)
1.1-2.1
0.0- 0.7 8.4-22.4
244
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As can be seen from this table, intermittent spreading
techniques maintained aerobic conditions in the soil, making
possible the oxidation of ammonia and organic nitrogen to nitrate.
Under continuous spreading, anaerobic conditions prevailed, and
ammonia was still present in significant concentrations at a
7- to 10-ft depth.
When a low-rate application system is used, the amount of
nitrogen applied to soil with sewage effluent is not much more
than can be removed by crops, according to Bouwer (133). A low-
rate system involving application to wheat at 2.5 and 5 cm/week
was cited. At the lower application rate, 92 percent of the
nitrogen in the wastewater was removed in the soil, while at the
5 cm/week rate, 60 percent was removed. On the other hand, when
animal waste slurries or other wastewater with a relatively high
nitrogen content are applied, the amount of nitrogen supplied
may far exceed that which can be utilized by crops. To obtain
significant nitrogen removal under these circumstances, Bouwer
suggests that the system be designed to stimulate denitrification
in the soil. He cites an instance where this was done by instal-
ling an artificial barrier to water movement at a depth of 2 m,
causing the formation of an anaerobic region. Ammonia and organic
nitrogen in applied wastewater were converted to nitrate in the
upper, aerobic region of the soil, which was then denitrified
in the lower, anaerobic zone. The system removed 96 to 99 percent
of the total nitrogen applied at rates of 1 to 2 cm/day.
A study (1233) conducted at the sewage farm in Arroyo Grande,
California, indicated the pulse like effect of intermittent
spreading on the nitrate level in the percolate. During weekly
flooding of the field with settled sewage effluent, the upper
foot of the soil adsorbed 2,840 Ib of organic and ammonia nitro-
gen/ac. As the fields drained and were exposed to oxygen, this
adsorbed nitrogen was rapidly converted to nitrate, resulting
in nitrate levels of 1,000 to 2,000 mq/t in the soil solution
of the top foot. Consequently, the subsurface drainage water
exhibited pulses of high nitrate concentration when flooding
began. After all the nitrate was leached from the topsoil by
the applied wastewater, the nitrate content of the drainage water
returned to a low level.
When the South Tahoe Public Utilities Department sprayed
treated sewage on forested hillsides in the fall of 1963,
nitrogen removals by the soil mantle were more than 65 percent.
The removals dropped to 26 percent in the winter when the ground
was frozen. Significant amounts of ammonium ion were present
in the upper 4 in of soil, but nitrate levels were low at all
depths. The removal of nitrogen was attributed to the denitrifi-
cation in the soil mantle under anaerobic conditions (1233).
245
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The California State Health Department (1233) reviewed
nitrogen removals and management at a number of well-injection
waste disposal/recharge systems. Results of a six-month study
on the injection of slow sand filter effluent are shown
in Table 91 .
TABLE 91 . NITROGEN TRANSFORMATION IN RECHARGE
AQUIFER. MG/l (1233)
Organic-N Ammonia-N Nitrite-N Nitrate-N
Natural water in
the aquifer .4-.5 .2 0.0 .4-4.8
Filter effluent
before injection 2.2 1.5 .01 21.3
Recharged water in
aquifer, 20 ft
from injection well 1.4 1.2 .21 18.2
Recharged water
mixture in aquifer,
500 ft from injec-
tion well .9 .8 .003 6.1
Analysi s
of nitrate in
tion, as well
nitrogen level
a distance of 500
of data in Table 91 indicated that the decay
the aquifer was due to some extent to denitrifica-
as dilution. Considering such decay, the nitrate
was expected to fall to 10 mg/l before traveling
ft in the aquifer. Further tests revealed an
anaerobic, microbiologically active zone in the aquifer in the
vicinity of the injection well. During the tests, nitrate was
largely removed by microbial denitrification within 150 ft of
the injection wel1.
In summary, data indicate that, under proper management
conditions, land application of wastewater effluent offers the
potential to efficiently remove nitrogen from wastewater and
protect groundwater from nitrate contamination. The most
successful programs stressed an appropriate flooding/drying
schedule to promote both aerobic nitrification and anaerobic
denitrification processes, in order to ultimately convert
ammonia nitrogen in the wastewater to nitrogen gas. However,
if not properly managed, a definite danger exists of polluting
groundwater resources with excess nitrates.
246
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A number of investigators have mentioned the possibility
of groundwater contamination by nitrogen forms from land
application of sludges (377, 690, 705). Yet few quantitative
studies were found. Walker (1440) concluded that trenching
was a viable method to dispose of sludge with minimal ground-
water contamination by nitrates. Results of his study showed
no increase in pollutants (with the exception of chlorides) in
groundwater monitoring wells for up to 19 months after sludge
entrenchment. There was evidence that nitrate pollution would
become a problem when sludge dried out and became aerobic. In
order to prevent excessive nitrogen from reaching groundwater,
it was recommended that the site be underdrained, and that the
drained water be used to irrigate surrounding cropland.
Five sites receiving applications of municipal sewage
sludge were examined by Prothero (1124). At all sites, the
ammonia nitrogen and total nitrogen concentrations of the
treated soils compared closely to the control soils at a depth
of 50 cm in the soil profile. The nitrate concentrations at
this depth were above those of the control soils and increased
with increasing application rate.
Brown (151) discussed sludge application to croplands.
He stated that serious nitrate pollution becomes a potential
problem only when the plant-available nitrogen application to
soils exceeds the sum of gaseous nitrogen losses (through
denitrification and volatilization) and the nitrogen require-
ments of the crop. He divided the nitrogen content of sludge
into the soluble ammonia fraction (which is immediately avail-
able for plant uptake) and the organic nitrogen fraction (which
must be converted to inorganic nitrogen by soil microorganisms
before it is available). The conversion of organic nitrogen
is estimated to occur at a rate of 10 percent the first year
that the sludge is applied and 5 percent per year in the
following years. The 90 percent of organic nitrogen that is
not converted the first year is incorporated into humus-like
substances. As this organic nitrogen accumulates in the soil,
the rate of sludge application must be decreased if the balance
is to be obtained between yearly inputs and losses of plant-
available nitrogen to and from the soil.
Brown cited a study which shows that the nitrate content
of drainage waters from a sludge disposal site will be excessive-
ly high if no restrictions are placed on the quantity of nitro-
genous material applied to the soil. When 10 in/yr of digested
sludge were applied to soil lysimeters, the drainage waters
contained up to 449 mg/i N.
247
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Since 1971, the Metropolitan Sanitary District of Greater
Chicago has been monitoring a digested sludge land application
site. A 1974 report (1556) states that increases in nitrite
and nitrate concentrations have not occurred in groundwater
observation wells 1n areas receiving the sludge, indicating
that chese constituents are not migrating into the aquifer.
Recent studies of land application of wastewater effluent
indicate that the soil system is highly efficient in removing
phosphates from wastewater. Phosphate removal is both a function
of soil composition and travel distance. In most soils, phos-
phorus not taken up by plants is immobilized due to the adsorp-
tion of phosphate onto the soil. Adsorption is followed by fix-
ation by iron and aluminum oxides if the soil is acid, or by
precipitation into various forms of calciurn- phosphate if the
soil is basic (133). These reaction products are sufficiently
insoluble, so that phosphorus is held in the upper few centime-
ters of most soils, and very little phosphorus moves into the
groundwater (798). However, in the case of acidic, sandy soils
with no iron or aluminum oxides, little phosphate is fixed. Thus,
it may be necessary to remove phosphorus from wastewater before
its application to such soils (133).
Hook et al. (595) Deported that under proper management,
most of the phosphorus n wastewater remains in the soil at the
disposal site or leaves as a nutrient in harvested crops. They
found that soils differed in their abilities to retain phosphorus
In a heavy-textured soil high in iron and aluminum oxides and
hydroxides (sesquioxides) , phosphorus from effluent irrigation
did not increase in the soil below a depth of 1 ft after 7 yr of
irrigation. In a light-textured soil with half as much sesqui-
oxides, phosphorus content increased to a depth of 3 ft after
6 yr of treatment.
Phosphorus removals at Flushing Meadows, Arizona, (134)
were found to be basically dependent upon the distance traveled
by the wastewater through the soil. The chief removal mechanism
was precipitation as calcium phosphate or magnesium ammonium
phosphate, since the soils tested contained little iron,
aluminum oxides, or other phosphate fixing materials. Under-
ground travel distances of 30 ft produced 50 percent reduction;
distances of several hundred feet were found to be sufficient
for 90 percent phosphorus removal. However, the capacity of
the soil to remove phosphorus decreased after the start of
the project, holding stable at approximately 50 percent removal.
The fact that soils gradually lose their capacity to adsorb
phosphates over long-term application is substantiated by
Barrow's detailed analysis of this phenomenon (68). He
concluded that previously applied phosphate had been converted
to a form that was occupying phosphate adsorption sites, thus
reducing the capacity of the soil to further adsorb phosphate.
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Continuing studies at Lake George, New York (47), have
shown that significant amounts of nitrates appear to reach
the waters tributary to the lake. However, the wastewater land
treatment system appears to remove essentially all phosphorus,
thus reducing the potential for algal bloom in the lake.
Dugan et al. (352), in their work on land disposal of
wastewater in Hawaii, noted that phosphorus removals of over
95 percent within a 5-ft depth of percolation were obtained
when secondary wastewater was applied to grassed areas.
Suspended solids are removed very effectively by land
application systems. In fact, one problem encountered in the
spreading of wastewater is that nearly all suspended solids
are filtered out in the top few inches of soil. This can cause
clogging of soil pores and reduction of infiltration rates.
There is no danger of groundwater contamination in applied
wastewater from suspended solids. Literature data on suspended
solids are not extensive, because experience has shown land
application to be capable of removing virturally all solids at
the surface. A study at Whittier Narrows, California (1233),
showed suspended solids removal, due to percolation, of 95
percent. At Flushing Meadows (134), the suspended solids concen-
tration of the percolate was essentially zero, even though the
solids in the wastewater applied reached 100 mg/l. Similar
findings were reported at Lake George (47), where virtually
all BOD and suspended solids were removed from percolated
effluent.
Results from a spray irrigation and runoff system used to
dispose of a cannery waste (87) showed that even with a runoff-
type system, suspended solids removals averaged 97 percent.
Dugan et al. (352) reported similar high suspended solids
removals with application of secondary effluent to Bermuda
grass in Hawaii.
ELEMENTAL CONTAMINANTS
Municipal wastewater and sludge contain small amounts of
nearly all metals. The degree to which a particular soil will
protect underlying groundwater through removal of contaminants
is primarily determined by the chemical and physical composition
of the soil. Removal can occur through such processes as precip-
itation of solid phases, ion exchange, and adsorption. These
processes are in turn controlled by soil pH, the oxidation/
reduction potential, clay content, the presence and type of
organic material, and the extent of soil saturation.
The general nature of reactions of sewage wastes with soil
is well known. With time, wastes applied to land are broken
down, and the dissolved constituents become part of the soil
solution. Released cations can exchange with those already
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on exchange sites in the soil. Metals as ions or in the
colloidal state can be adsorbed onto soil surfaces. When the
levels of ions in solution exceed the solubility of correspond-
ing solid phase compounds and minerals, those compounds can
precipitate. When the solubility of solid phase compounds and
minerals exceeds the levels of corresponding ions in solution,
the compounds can dissolve. Constituents are also ingested
by soil microorganisms and incorporated into soil organic matter.
Ions that are not removed by any of these processes but
that remain in the soil solution are available for uptake by
plant roots or leaching by water moving through the soil profile.
Lindsay (798) studied the composition of the soil solution,
concluding that it is controlled by the solubilities of solid
phases. Thus, precipitation and dissolution reactions determine
the activity of ions in solution, which in turn governs ion
exchange.
Lindsay (798) also recognized the importance of the forma-
tion of metal-organic complexes and chelates in increasing the
solubility and mobility of metals in soils. Brown (151) stressed
this point. He found that it is misleading to predict the
availability of metals in soils from their solubility in distilled
water. He cites evidence that plant uptake of metal ions could
not be predicted by the water solubility of the solid compounds.
This may be because the soil solution in the vicinity of roots,
unlike distilled water, is mildly acidic and contains organic
metal-complexing agents (798).
All of the trace elements for which water quality criteria
have been established may occur as either soluble or insoluble
metal-organic complexes (1323). Low molecular weight organic
molecules tend to increase the penetration of complexed metal
ions into the soil, while high molecular weight molecules and
their complexed ions may be filtered out by the soil (133).
The chemistry of the metal organic complexes is complex, and
present knowledge of organic forms of the elements is insuffi-
cient to generalize. Where concentrations of trace elements in
soil solutions are in excess of those predicted from inorganic
solubility product considerations, the element is thought to
occur in organic form. Because of this lack of knowledge, the
literature discusses mainly the chemistry of inorganic forms.
This is unfortunate, because most of the metal ions in waste-
water and sludge probably occur in complexed form.
Most metals are less mobile under aerobic and basic soil
conditions than they are under anaerobic and acid conditions
(133). The neutral water extract of soils contains less heavy
metals than acidified extracts of the same soils, indicating
the importance of the pH factor in influencing the mobility of
metals (1173). Lindsay (798) cites evidence that for zinc and
copper there is a 100-fold increase in ionic activity for each
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unit decrease in soil pH. Normally insoluble metals may
become mobilized in the event of a change in the characteristics
of percolating water, such that acidic or anaerobic conditions
occur.
The removal of metals by ion exchange or adsorption depends
upon the availability of exchange and adsorption sites in the
soil as well as on the factors (ionic activity and pH) just
discussed. Therefore, the clay content of the soil is important;
clay soils provide more exchange and adsorption capacity than
sands and gravels (1173). A correlation also exists between
the form in which the elements occur in solution and their
removal by exchange or adsorption. In general, elements that
occur in solution as anions or neutral molecules pass through
soils more readily than do elements that occur as cations.
Inorganic arsenic, selenium, and fluorine in aerated soils
occur as anions or neutral molecules. Although there are
exceptions (depending upon the chemistry of the system),
inorganic cadmium, copper, chromium, lead, mercury, silver, and
zinc most commonly occur in fresh waters and soil solutions in
the inorganic form as cations (1323).
Many conclusions in the literature lack substantiating
experimental data but are based on a knowledge of chemistry.
The use of sanitary sludge for land reclamation projects may
increase the toxic metal content of the soil. The work of
Lejcher and Kunkle (776) showed significant reductions in the
leachate of iron, aluminum, copper, manganese, and sulfate
combined with increases in cadmium, zinc, and chromium when
304 tons dry wt/ac were applied. In this experiment, the
resulting pH of the surface soils was 6.2, still low enough to
account for the mobilization of the metals. Higher application
rates or the utilization of sludges with higher buffering
capacities may help immobilize the metals that showed increases.
The work by Sopper et al. (1299) revealed similar results.
Bernard (106) considered metal concentrations in sludge
a prime deterrent to landfilling as a disposal method, but he
concluded that metals tend to concentrate in the upper soil
layer rather than leaching through the soil profile. Limited
monitoring by Lofy (ongoing research) at eight landfill sites
accepting wastewater sludge indicated that there was measurable
migration of certain metals to considerable distances from the
disposal area. Lead was particularly mobile. Preliminary
indications are that many metals, such as chromium, are strongly
attenuated by the majority of soils studied.
Prothero (1124) studied five sites where sludge is disposed
of on land. The heaviest application rate at a site was 500 dry
metric tons of sludge/ha. Water samples were collected from
wells in close proximity to the sludge disposal locations and
were analyzed for cadmium, chromium, manganese, nickel, and
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zinc. The trace element content of the applied sludge fell
within ranges generally reported for municipal sludges, except
for the chromium content of one sludge (17,700 ppm). Results
showed the cadmium, chromium, and nickel contents of the well
waters were generally undetectable while the copper, manganese,
and zinc concentrations did not exceed 0.3, 0.01, and 0.04 ppm,
respectively. There was an exceptionally high manganese content
(1.2 ppm) at one site, however, which was apparently due to
strong reducing conditions. The other well samples were below
maximum allowable drinking water concentrations set by the
Public Health Service for each metal.
Walker (1440) considered groundwater contamination by
heavy metals at a sludge entrenchment- site. For 19 months after
sludge was buried at 350 to 500 tons/ac dry sludge solids, no
metals were detected in groundwater observation wells.
Data from actual land application operations, which
included arsenic and selenium determinations, were not found.
Predictions of removal of these metals based on chemical
properties are also hindered by limited understanding of the
reactions involved.
Only one reference was located that provided experimental
data on cadmium in relation to wastewater applied to land.
At Flushing Meadows, Bouwer et al. (134) found that cadmium in
wastewater applied to land in shallow basins showed very little
change due to migration through the soil. The cadmium concen-
tration dropped only slightly from 7.7 yg/£ to 7.2 yg/£.
Aerobic conditions and alkaline pH prevailed in the soil studied.
A study (1323) by the California State Water Resources Control
Board (CSWRCB) reviewed available data on soil reduction of
cadmium in effluents, concluding that information is not
sufficiently detailed to allow adequate evaluation of cadmium
concentrations in water reaching groundwater basins. According
to the study, it must be demonstrated that a particular soil
is able to reduce the concentration of cadmium to a level that
is acceptable for drinking water. Otherwise, the study advised
against the use of wastewater effluents with concentrations
above this limit for groundwater recharge operations involving
percolation through soil.
Bouwer's work at Flushing Meadows (134) found that the
copper concentration of applied wastewater was reduced about
86 percent by passage through the soil. This removal occurred
rapidly, usually in the first 30 ft of downward flow.
f
Iron and manganese in we!1-oxidized soils are characterized
by the formation of highly insoluble oxides and hydroxides.
However, at low pH and under reducing conditions, these metals
can be solubilized and become mobile in the soil as Fe2+ and
Mn2+. Amramy (26) conducted a study of sewage lagoon effluent
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spreading on sand dunes. He found that after a subsurface
travel distance of 8 m, the concentration of iron in the waste-
water actually increased from 0.28 mg/£ to 0.57 mg/i. The
manganese concentration increased from 0.08 mg/i to 0.19 mg/i
after 25 ft of travel through sand. Wesner and Baier (1484)
found a similar phenomenon when tracing the underground move-
ment of wastewater after injection. Over a travel distance of
400 or 500 ft, the concentration of iron did not change.
Manganese concentrations, on the other hand, increased up to
300 percent in the first 100 ft of subsurface travel. Anaerobic
conditions, favoring formation of soluble manganese compounds,
are mentioned as the possible cause of the large increase.
These conditions are most likely caused by biological oxidation
depleting oxygen near the point of injection, thus causing the
reduction of manganese to soluble forms. However, after greater
travel distances and the return to aerobic conditions, the
manganese may revert to insoluble compounds and be removed from
the migrating water.
Ragone et al. (1129) reported on deep well recharge expe-
riments conducted in Nassau County, Long Island. Tertiary
effluent was recharged by a deep well into the Magothy aquifer,
the primary water supply source for Nassau County. As of
September 1972, 12 recharge tests had been run since the incep-
tion of the recharge program in September 1968. Although
the iron concentrations of reclaimed and native water averaged
0.44 mg/i and 0.24 mg/i, respectively, the iron concentration of
the mixed (native and reclaimed) water at times exceeded 3 mg/i .
The authors mentioned several sources that could account for
the increase in iron concentration, but the most probable source
was the pyrite native to the Magothy aquifer. During recharge,
the natural reducing condition in the aquifer was replaced by
a progressively more oxidizing environment. The initial
response to this change was the oxidation of pyrite, which
released Fe+z, $04-2, and H+ into solution. Eventually, ferric
hydroxide precipitated, and the Fe + 2 concentration decreased.
The exact oxidation mechanism apparently involved inorganic
and/or organic constituents in the reclaimed water, because
water from the public potable water supply system caused no
increase in iron concentration when injected into the aquifer.
The Flushing Meadows study by Bouwer et al. (134) contained'
data on the reduction of lead by soil infiltration. Bouwer found
that the wastewater concentration of lead decreased by 20 percent
after significant travel distance underground. Apparently, a
small portion of the lead was tied up rapidly (within 50 ft),
while the majority was unaffected by further travel. It should
be noted, however, that the soil examined in these experiments
was a sand with limited adsorption and exchange capacity for
many trace elements.
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Both mercury and zinc form insoluble compounds in soil,
lowering the activity of the ions in solution so that little
movement occurs. Mercury is particularly insoluble as phosphate,
carbonate, or sulfide. However, under low pH conditions, the
metals may become mobilized. They can also form soluble
complexes that affect their mobility under certain circumstances
(798). At Flushing Meadows (134), underground percolation
through 100 ft of sand produced 40 percent removal of mercury
and approximately 58 percent removal of zinc. Further travel
produced no further removal of mercury but reduced the zinc
concentration to about 20 percent of that in the applied waste-
water. The final concentrations of mercury and zinc were
.00014 mg/£ and .037 mg/£, respectively.
In conclusion, the information garnered in this study of
the literature is inadequate to define fully the chemical
behavior of elemental contaminants in the soil and their fate
as they percolate through the unsaturated zone. The CSWRCB
recommended, on the basis of its literature review (1323), that
wastes containing concentrations of certain metals above those
acceptable for drinking water supplies should not be applied to
land, unless it can be demonstrated that contamination of
groundwater does not occur.
BIOCIDAL CONTAMINANTS
Use of chemical pesticides in agriculture Generated many
studies of the potential harmful effects of these compounds
on land, crops, surface water, and groundwater. However, there
is little data available on biocides in municipal wastewater or
on the potential dangers of groundwater contamination through
land wastewater application operations. This lack of informa-
tion is understandable considering the minimal role that munici-
pal wastewater plays in transporting biocides to the land.
Biocides come in contact with the land through a number of
activities, primarily by direct application to the land for
pest control. Return irrigation water, and spills and wastes
from pesticide manufacturing operations also bring biocides in
contact with the land.
The reactions of pesticides with soil has received limited
attention in the literature. Volatilization, chemical degrada-
tion, and absorption by plant roots and seeds apparently remove
a small portion of pesticides reaching the soil. A more signi-
ficant process for pesticide removal may be microbiological
degradation. Although this process is very slow in some cases,
often taking several years, microbiological degradation accounts
for the breakdown of a remarkable variety of organic compounds.
Pesticides that are not removed from the soil column or broken
down by these processes may be available for leaching into
groundwaters.
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Gerakis and Sficas (464) reviewed the literature on pesti-
cide degradation and leaching. They reported that the most
important factors involved in these two processes are soil
temperature and moisture, organic matter and clay content, soil
management practices, pH, and species and population density of
microorganisms present. The presence of organic matter and clay
in the soil appears to be positively correlated with adsorption
of pesticides onto soil particles.
Van Bladel and Moreale (1417) studied herbicide adsorption
onto clay minerals. They found that adsorption increased with
the polarizing power of the exchangeable cation, and concluded
that adsorption appears to be one of the most important factors
in reducing pesticide removal from soil layers by leaching.
Laboratory studies of land-applied sludge were conducted by
the University of Nebraska (1474) to determine degradation rates
in groundwater for selected herbicides. The study indicates that
herbicide degradation was much slower in groundwater than in soil.
O'Connor and Anderson (1014) analyzed factors affecting adsorption
of the herbicide 2,4, 5-T on four soils in the western United
States. The study found that organic matter contributed to
adsorption, while oxides of iron and aluminum did not.
Gerakis and Sficas (464) cite evidence that pesticides
differ in their mobilities in the same soil. One study showed
that (1) acidic compounds are relatively mobile, (2) phenyl
ureas and triazines are of intermediate to low mobility, and
(3) organochlorine compounds and organic cations are least
mobile. Further data that were reviewed supported the conclu-
sion that under normal agricultural practices and rainfall, it
is very unlikely that pesticides may be leached deeply enough
and in such quantities as to cause appreciable contamination
of groundwaters.
A California State Water Resources Control Board report
(1323) mentioned that pesticides are adsorbed by soil clays,
iron aluminum oxides, and especially by organic colloids, and
that they are susceptible to microbial decomposition. However,
the amount of biocides in average municipal wastewater and
sludges, was found to be so minimal that the spreading of
municipal waste on land offers extremely low potential for
groundwater biocide contamination.
SYNTHETIC/ORGANIC CONTAMINANTS
One of the most intensely debated questions regarding land
application for treatment and/or disposal of municipal wastes
concerns the problem of residual organic contaminants. Refrac-
tory organic compounds may survive conventional treatment
processes and penetrate through the soil to contaminate ground-
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water supplies. The controversy centers around the fate of
residual organlcs within the soil systems, including such issues
as the synerglstlc effects between organlcs and Inorganics or
other groundwater and soil constituents, or conversion of safe
organlcs to hazardous compounds 1n the soil. Despite this
controversy, no literature was found concerning groundwater
pollution by the synthetic/organic contaminants in municipal
wastes as a result of land application.
The absence of literature concerning the movement through
soil of synthetic/organic contaminants 1n applied sludge 1s
not surprising, since the specific organic makeup of sludge 1s
unknown. Some of the chemicals of concern (PCB's, polycyclic
aromatics, and other chlorinated hydrocarbons, etc.) have low
solubilities in water 1n comparison with vapor pressures. As
a result, there 1s a distinct possibility of vaporization when
sludges containing these chemicals are applied to the land surface
The California State Water Resources Control Board (1323)
cited a study carried out 1n Colorado that compared the nature
of the soluble organic material in the soil profiles under a
feedlot and under grassland with selected ground, well, and
river waters. It was concluded that the major portion of the
soluble materials in all the waters was polymeric. The soluble
organlcs under grassland were essentially the same as those
under the feedlot, although phenols were present in greater
abundance in the manure and surface soil of the feedlots. About
13 percent of the soluble material in the soil profiles was
carbohydrate (polysaccharides), and much of the remainder, based
on IR spectra and reductive degradation procedures, appeared to
be polymerized aromatic structures. This report by the CSWRCB
Interpreted these observations to indicate that the soluble
organlcs under wastewater-treated soils would be similar to
those under feedlot manure or grassland.
BIOLOGICAL CONTAMINANTS
Most available data suggest that virus, bacteria, and
other biological pathogens present in wastewater and sludge
are removed or Inactivated by percolation through soil.
The California State Water Resources Control Board study
(1323) provides a summary of the fate of viruses, bacteria,
protozoa, and parasitic worms in wastewaters applied to land.
The summary states that most of these pathogens prefer warm-
blooded animals as their habitat and do not flourish in the
soil environment. When introduced into soils, the pathogens
do not compete well with the vast number and variety of normal
soil inhabitants and are subject to attack by antagonistic soil
species. The time necessary for their ultimate destruction
varies, according to species and environmental conditions.
A compilation of pathogen survival data in the literature is
shown in Table 92 below.
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TABLE 92 . SURVIVAL OF PATHOGENS
IN SOILS (1323)
Ascari s 1 umbri coides ova
Endamoeba histolytica cysts
Salmonella species
Coliform group organisms
Q-fever organisms
Bruce!la abqrtus
Tuberculosis bacteria
Enteroviruses
2.5-7 years
8 days
6 hours
133 - 147 days
148 days
30 - 100 days
6 months
12 days
The most persistent pathogens in soils appear to be ova,
cysts, and spore-forming bacteria. The survival of enteric
viruses in soil has not been thoroughly studied. The dependence
of enteric viruses on specific host organisms for reproduction
suggests that they would not multiply and would not be expected
to survive for a long period of time, although survival may,
at times, be long enough to cause public health concern.
Two and a half years of continuous observation was conducted
of wastewater reclamation by 1andspreading in Lodi , California
(1175). It was found that the MPN of coliform group organisms,
which averaged 1.9 x 1Q8/100 ml in the wastewater, was consis-
tently reduced to less than 1/100 ml after 4 to 7 ft of soil travel.
The average percolation rate was 0.3 ft/day in coarse-textured
Hanford sandy loam. It was observed that the number of coliforms
penetrating 1 ft or more was essentially independent of the
coliform concentration of the wastewater.
At Whittier Narrows (892), percolation tests showed that
vertical percolation of wastewater through 4 to 7 ft of soil
is an effective method of removing bacteria of fecal origin,
despite heavy growth of coliforms of soil origin. The formation
of an organic-microbial slime layer at the water-soil interface
was found to increase the efficiency of the filtering action.
Lejcher and Kunkle (776) reported, with regard to biological
pathogens, that no fecal coliform bacteria were found in the
runoff from the sludge-treated plots (treated at a rate of
78 to 304 dry tons of sludge/ha) during the year immediately
following application. They concluded that it appeared unlikely'
that pathogens would survive in the runoff. Limited transfer
of microbial cells through the soil profile and rapid die-off
of the pathogens in the soil matrix suggested low-risk
conditions (619, 917).
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Research by Bernard (106) on the disposal of sewage sludge
to sanitary landfills concluded that pathogens tend to concen-
trate in the upper soil layer, rather than leaching through
the soil profile. In contrast, Lofy (ongoing research) found
that fecal coliform and fecal streptococcus are particularly
mobile and can travel considerable distances.
Results from studies at Flushing Meadows, Arizona (134)»
show that fecal coliform density was reduced significantly in
the first 2 or 3 ft of travel. Buuwer found that fecal coliform
density at a particular depth tended to decrease with increased
flooding time. The peak bacteria density invariably appeared
immediately after flooding was resumed. The concentration of
fecal coliforms was consistently decreased to less than 10/100 mi
after 100 ft, and to 0/100 ml after .300 ft of travel.
A project at Santee, California (1215), is famous for its
pioneering work in the reclamation of domestic sewage for
recreational lakes. Travel of secondary effluent through
1,500 ft of very coarse sand was sufficient to remove all fecal
coliforms. Sampling showed that most of the coliforms were
removed in the first 200 ft.
At Orange County, California (1484), tests conducted on
a well injection system showed coliform organisms 30 m from
the injection well, but none approximately 80 m from the well.
The results indicated that fecal conforms are more easily
removed by underground travel than other coliforms. Some of the
other coliforms may have been supported by nutrients 1n
the effluent.
Results of percolation tests at Lake George, New York
(1129), showed again that: percolation of applied secondary
effluent through 5 to 10 ft of soil in two different beds was
sufficient to remove essentially all coliform organisms.
Browning and Mankin (155) reported an unusual case of
disease outbreak due to contamination of groundwater well
supplies by land application of treated sewage. In Madera,
California, undisinfected secondary effluent was used to irrigate
a pasture located adjacent to a deep well drawing part of the
city water supply. The wastewater migrated through gopher holes,
filling a construction pit around the well, and eventually
flowed into the well itself.
On the basis of experience and results of full-scale,
long-term wastewater reclamation studies, the CSWRCB (1323)
concluded that, although soil is an excellent media for removing
bacteria, a small fraction of the fecal coliform bacteria there-
in may reach qroundwater reservoirs at hiqh percolation rates.
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Horizontal travel of viable fecal coliform bacteria in the
aquifer does not appear to occur to a significant degree. The
available data on horizontal travel, however, are inconclusive.
Further investigation of the transport and survival of pathogenic
bacteria in groundwater, therefore, is required.
Research shows that travel through soil removes significant
amounts of viruses, primarily through adsorption. Adsorption
is influenced by the pH and ionic strength of the soil solution.
The available information indicates that adsorption of virus
by soil is nearly complete at pH 7 or less, but decreases as
the pH value increases above 7. This is mainly because the
overall electric charge surrounding both the virus and soil
particles becomes increasingly negative as pH levels increase
and, therefore, mutual repulsion occurs (1323).
It also appears that increasing cationic strength of the
percolating water or soil solution increases virus removal. At
the pH values normally encountered in wastewater, viruses are
slightly negatively charged. The presence of calcium, magnesium,
sodium, aluminum, and other positive ions in the soil solution
decreases the potential for negatively charged soil and virus
particles to repel each other. This results in the formation
of soil-cation-virus bridges that immobilize virions (1323).
The ionic strength in percolating wastewater is usually
sufficient so that it does not limit adsorption. In circum-
stances where ionic strength is significantly decreased,
however, desorption of adsorbed viruses may occur. Organic
matter in wastewater can also compete with viruses for adsorp-
tion sites. In laboratory studies, when virus-adsorbed clay
particles were washed with distilled water, an essentially
complete desorption and reactivation of viruses took place.
In field conditions, other mechanisms in soil systems may
inactivate or destroy adsorbed viruses before they are subject
to desorption (1323). In addition to pH and ionic strength,
the clay and organic matter content of soil evidently influences
adsorption to some degree. In general, soils of higher clay
and/or organic matter content are more effective in adsorbing
viruses (1323).
Definitive work on virus interaction with soil was conducted
at Santee, California, where extensive studies showed that
percolation through several hundred feet of soil consistently
removed all virus from secondary effluent (1215).
Other studies also supported the conclusion that soil
effectively removes viruses. Viral analyses in Hawaii by Dugan
et al. (352) showed that test soils in 5-ft lysimeters were
completely effective in removing viruses. Brief tests at
Whittier Narrows, California (892), achieved complete removals
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of Sabin Type III polio-virus vaccine. Although 250 plaque-
forming units (PFU) of enteric viruses/^ were present in the
applied wastewater, no measurable concentrations were found below
2 ft in the percolate.
In 1974 at Flushing Meadows (134), virus analyses were
performed bimonthly to determine the fate of viruses in the
soil system. Secondary effluent was allowed to infiltrate into
six parallel horizontal basins consisting of 60 to 90 cm of
fine loamy sand underlain by several coarse sand and gravel
layers to a depth of 75 m, where a clay layer begins. Observa-
tion wells were installed in line across the basin area. No
viruses were detected in any of the wells at any time during
each flooding period. Gilbert et al. (475) stated that the
failure to detect viruses in the wells indicates that the virus
count was reduced by at least 99.99 percent within 3 to 9 m
of basin soil.
Romero (1175) reviewed the studies performed ^or the
Department of the Army on sands ranging in classification from
silty sand to coarse-granite alluvium. Results indicated that
the bacterophages Tl, T2, and 65 are more effectively retained
in the finer sands, particularly in those containing a relatively
high percentage of clay and silt. Virus removal was shown to
increase with decreasing particle size. The greatest percentage
of removal took place in the uppermost portion of the sand
columns tested." It was shown that for a well-sorted sand of
particle size averaging 0.12 mm, the removal efficiency in 2 ft
of penetration was 99.999 percent.
Young and Burbank (1548) described studies of virus removal ;'
in Hawaiian soils. In the laboratory, columns of various types
of Hawaiian soil were subjected intermittently to percolating
water with a known concentration of virus (coliphage T4B II mutant,
and polio virus Type II (Lansing) H8), simulating the action of
cesspool leaching. The effluent from each soil column was
analyzed for viral content. Coliphage T4 was applied to
slightly acid soils (pH 5-6) at a concentration of 2.5 x 10b/m£.
Percolation through 2.5 to 6 in of soil was 100 percent effec-
tive in retention of the virus. Slightly alkaline soil was
less effective, removing only 67 percent of applied coliphage
and 35 percent of applied polio virus in 15 in. Removal of
polio virus Type II was less complete; 6-in columns were able
to effect only 99.3 percent removal with an initial feed concen-
tration o.f 1.5 x 105 pfu/m£.
Wellings et al. (1479) found that virus can be isolated at
the 6.5-m level below a spray irrigation' field. Another study
by Wellings et al. (1477) measured virus migration through the
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ground from chlorinated packaged plant effluent applied to a
cypress dome. Both horizontal and vertical migration was
detected at distances of approximately 7 m for polio and
coxsackle viruses. Wells beyond that distance showed no virus.
The survival of virus within the dome was at least 28 days.
At Fort Devins, Massachusetts, where a land application
site has been 1n operation for over 30 years, Schaub et al.
(1210) studied the removal of bacteria from unchlorlnated
primary effluent applied to soil cells. Using tracer f2 bac-
terlophage and the enterovlruses polio virus I and EMC virus,
it was demonstrated that tracer bacteriophage penetrated Into
the groundwater along with the percolating wastewater. The
concentration 1n the groundwater stabilized at almost 50 percent
of the applied virus concentration. The tracer and entero-
vlruses were sporadically detected at horizontal distances up
to 600 ft from the application point.
Lance et al. (750) passed secondary sewage effluent con-
taining 3 x 104 pfu/m£ polio-virus Type I (LSc) through' 250 cm-
long columns packed with calcareous sand from an area 1n the
Salt River bed used for groundwater recharge of secondary sewage
effluent. Viruses were not detected in 1-ml samples extracted
from columns below the 160-cm level, but were detected 1n 5 of
43 IQQ-ml samples of the column drainage water. Most of the
viruses were adsorbed in the top 5 cm of soil. Virus removal
was not affected by the Infiltration rate, which varied between
15 and 55 cm/day. Flooding a column continuously for 27 days
did not saturate the top few centimeters of soil with viruses
and did not seem to affect virus movement. Flooding with
deionized water caused virus desorption from the soil and
increased virus movement through the columns. Drying the
soil for one day between applying the virus and flooding with
deionized water greatly reduced desorption, and drying for
five days totally prevented desorption. The investigators
concluded that large reductions (99.99 percent or more) of
virus are expected after passage of secondary effluent through
250 cm or more of calcareous sand, unless heavy rains fall with-
in one day after application of sewage.
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PATHWAYS TO MAN
AIR
INTRODUCTION
Health-impairing contaminants contained in wastewater are
most likely to return to man by way of surface or groundwater
migration; however, some contaminants may also be transmitted
through the atmosphere. Aerosolized particles ranging in size
from 1 to 40 microns are generated as an incidental by-product
of various wastewater treatment processes and may contain
health-impair ing contaminants that can be directly inhaled by
man. Glaser and Ledbetter (481) found that 40 percent of the
particles aerosolized during activated sludge treatment were
smaller than 10 microns-. Larger particles pose a minimal
hazard to respiratory systems but may be ingested; however,
aerosols of 10 microns or less will be captured in the upper
respiratory tract, transported through the pharynx, and will
enter the digestive tract.
Activated sludge, trickling filters, and spray irrigation
of wastewaters and liquid sludges are considered to be the
major potential sources of aerosols. In addition, vaporized
contaminants are emitted from sludge incinerators and furnaces
used for activated carbon regeneration.
Bubbles are produced in wastewater by natural biological
action or by mechanical aeration during activated sludge and
similar processes. These bubbles burst at the surface, eject-
ing contaminant-bearing water droplets into the air. When
wastewater is sprayed, as in land application, large numbers
of particles are released into the atmosphere. Aerosols con-
taining wastewater contaminants are also produced above bodies
of water into which wastewater effluent is released as a
result of wind spray, breaking waves, splashing rain, or
bursting bubbles.
Certain contaminants are floatable and tend to concen-
trate in a layer on the water surface. Particles that are
aerosolized from this layer may contain high concentrations
of these contaminants relative to their concentrations in the
bulk of the liquid. For example, Blanchard and Syzdik (125)
found the bacterial concentration of Serratia marcescens in
drops from bursting bubbles to be from 10 to 1000 times higher
than that of the tap water in which the bubbles formed. These
investigators suggested that DDT and several species or
organisms could be expected to concentrate in surface films
and be transferred to the atmosphere.
262
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Direct vaporization of certain contaminants also occurs.
For instance, low solubility contaminants such as pesticides,
polychlorinated biphenyls, or hydrocarbons have high evapora-
tion rates from water bodies to the atmosphere, according to
a paper cited by Mackay and Leinonen (827).
The viability of biological aerosols from sewage treatment
plants is affected by distance traveled, deposition, dispersion,
and hostile environmental conditions.
One study revealed the correlation between human respira-
tory difficulty and high concentrations of plankton in the
sea. This study has often been cited (125) as indicative of
the airborne, onshore movement of marine organisms. The
"blowback" of microorganisms rising to the surface from sub-
marine sewage outfalls may also have a significant, localized
impact on public health.
A considerable body of information has been developed in
medical literature pertaining to the movement of pathogenic
bacteria through air. In scientific literature topics studied
include atmospheric physics and dispersion modeling, aerosol
physics, liquid surface physics, dissolved air and froth
flotation technology, and other subjects indirectly related to
potential atmospheric pathways of contaminants from waste-
water treatment systems to man. Very little of this work,
however, deals directly with the subject of atmospheric path-
ways of contaminants from wastewater to man. Available
literature is listed in Table 93. Research in this area has
been principally concerned with the transport of biological
pa thogens .
ELEMENTAL CONTAMINANTS
For most elemental impurities, direct vaporization will
not occur. Few references were found in the literature re-
viewed that dealt directly with vaporization or aerosoliza-
tion from wastewater systems, although Haque and Freed (538)
included a discussion of the thermodynamics involved.
The atmospheric transport of mercury and dimethyl-mercury
has been studied and reviewed by Jernelov et al. (645), with
theoretical calculations performed by Baughman et al. (78).
On the basis of Baughman's calculations, it can be assumed
that the evaporative loss of these substances from aqueous
solutions may pose a problem under turbulent conditions.
Elemental mercury appears to be volatilized twice as rapidly as
the dimethyl form.
Soldano et al. (1295) measured organic and elemental
mercury concentrations in cities with central sewage facilities.
A broad range of mercury concentrations was detected in the
vicinity of treatment facilities. Numerous measurements were
in the range of 103 n g / m 3 , but values as low as 0.125 n g / m 3
263
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TABLE 93. LITERATURE REVIEWED PERTAINING
TO ATMOSPHERIC TRANSPORT
Contaminant Reference Number
Water Quality Parameters 481
Elemental Contaminants
Boron 1113
Mercury 78, 538, 645, 1295, 1302
Selenium 1114
Other (general ) 538
Biocidal Contaminants
Chlorinated 1532
hydrocarbons
DDT 553, 1472, 1551
Dieldrin 553, 1551
2,4D 553
Synthetic/Organic 1219, 1564
Contaminants
Biological Contaminants
Bacteria 79, 125, 238, 400, 481, 545, 568,
624, 649, 685, 765, 766, 792, 936,
983, 1069, 1073, 1099, 1100, 1234,
1302, 1304, 1463, 1560
Coliforms 2, 400, 568, 685, 983, 1099, 1100,
1234, 1304
Hepatitis virus 1151
Parasitic worms 1234
Protozoa 1234
Salmonella 568, 1234
Shigella 568, 1234
264
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TABLE 93 (continued)
Contaminant Reference Number
Virus 79, 238, 400, 649, 792, 1261, 1302,
1304
Other (general) 125, 238, 481, 568, 765, 766, 983,
1073, 1234
were also found. The atmospheric concentration of elemental
mercury fell off sharply with increasing distance from the
sewage plant, while the concentration of organic mercury rose
with increasing distance. The concentration maxima of organic
mercury at any particular plant was related to population size
of the area served. The authors concluded that sewage treat-
ment systems concentrate and re-emit elemental mercury after
some of it has been converted into the more volatile organic
forms by bacterial action.
BIOCIDAL CONTAMINANTS
Biocides tend to concentrate in the liquid surface and
aerosol fractions. However, the threat of adverse health
effects from aerosolized biocides contributed by sanitary waste
systems appears small in relation to that posed by other sources
of these contaminants. It is therefore not surprising that
only indirectly related research is available on the transport
of biocides from sanitary waste systems through the atmosphere.
Information on pesticides in rainfall and runoff is provided
by Weibel et al. (1472), and factors affecting the volatility
of DDT, dieldrin, and 2, 4-D from leaf and glass surfaces have
been examined by Hee et al. (553). Young and Heesen (1532)
conducted a study of the atmospheric transport of chlorinated
hydrocarbons to the waters of three southern California harbors.
They determined that atmospheric transport accounted for less
of the total concentrations of p,p'-DDT and PCB 1254 than
surface runoff.
SYNTHETIC/ORGANIC CONTAMINANTS
In general, insoluble or limited solubility organic
materials are concentrated at the air/water interface, in turn
producing a concentrated aerosol. The primary significance
of this surface concentration of organic materials and aerosols
lies in the effect that these materials may have on the sur-
vival of biological aerosols, Webb (1463) has conducted
265
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experiments showing enhanced survival of several species of
bacteria when amino acids, long chain protein degradates,
some sugars, and polyhydroxycyclohexanes were added to
bacterial suspensions before aerosolization. Hatch and
Wolochow (545) have also reviewed several aspects of this
problem, concluding that most of the organic compounds shown
to be the best protective agents for pathogen survival are
sugars or polyhydricalcohols.
BIOLOGICAL CONTAMINANTS
•A number of recent studies have examined the airborne
bacterial levels adjacent to treatment plant and land irriga-
tion sites. However, no correlation has yet been made between
specific atmospheric organism levels and the incidence of
disease (238). In their 1975 review of the literature,
Hickey and Reist (568) found only one survey that
had a direct bearing on increased disease rates associated
with occupational exposure to viable aerosols in wastewater
treatment operations. In this survey, the rate of pneumonia
incidence was identical for both water purification plant and
sewage plant workers, but the incidence of flu and colds was
higher among sewage workers by factors of 50 to 28 percent,
respectively.
The literature on airborne levels of micro-organisms is
difficult to interpret for several reasons. For instance,
sampling methods are not standardized, and a suitable indicator
organism is lacking. Also, there has been no research to
date suggesting that the traditional water quality indicator
organisms, the coliforms, provide a reliable indication of
pathogens present in aerosolized particles. In fact, the
half-life of airborne coliforms was found to be shorter than
that of the total bacteria in studies of the downwind disper-
sion of bacterial aerosols from activated sludge tanks performed
by Ledbetter and Randall (766). Pavoni and Tittlebaum (1069)
cited a study that found the survival rate of enterobacteriaceae
to be only 13 percent of that recorded for the total bacterial
aerosol emanating from an activated sludge unit. A second study
cited by Pavoni and Tittlebaum (1069) indicated that capsulated
organisms (klebsiella, aerobacter) are better able to survive
in the air than acapsulate organisms (escherichia) . The suit-
ability of coliforms and coliphages as animal virus indicators
was the subject of research by Fannin et al . (400). Evidence
collected from both trickling filter and activated sludge plants
showed coliforms to be far less stable than coliphages in the
airborne state. On the basis of this evidence, it was sug-
gested that coliphages may prove to be far more acceptable^
indicators of airborne animal viral contamination than coliforms
266
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Most research to date, however, has relied upon coliforms as pri-
mary indicator organisms to determine total bacterial concen-
trations. The bacterial concentrations at various distances
from treatment units, which are reported in the literature by
a number of investigators, are given in Table 94. This table
represents a sample of the available literature on activated
sludge and trickling filter plants.
The Adams and Spendlove paper (2) is unique in that it
reports recovery of airborne coliforms as far as 0.8 mi down-
wind from a large trickling filter plant. In contrast,
Hickey and Reist (568) cited one study in which no Escherichia
coli was found beyond 100 ft, and a second study that reported
no observable effects at greater than 150 ft downwind from a
trickling filter. Napolitano and Rowe (983) offer data show-
ing duplicate samples with coliform concentrations that differ
by as much as two orders of magnitude.
Relatively few investigations involved the recovery of
actual pathogens. According to research reviewed by Hickey
and Reist (568), of the total bacteria recovered near aeration
tanks, 10.5 percent were klebsiella, aerobacter or proteus
organisms. Despite repeated long-term sampling, neither the
shigella nor salmonella bacteria were recovered near the tanks,
a finding that was expected considering the small numbers of
these genera found in wastewater. Airborne Stapjry_J_o_coc_c_us aureus
hemolytic streptococcus, mycobarerium, and a cTcf-'fa s fTa c fl 1 i w e r e
also isolated.
A study conducted by Fanniri et al. (400) evaluated animal
virus concentrations. No animal viruses were found in aerosol
emissions from activated sludge and trickling filter plants
with sewage containing about 100 plaque-forming units (pfu)/-£.
Airborne coliphage viruses and coliform bacteria were re- -
covered at average levels of .23 to .3/m3 and 210 colonies/m ,
respectively. As shown in Table 94, the average coliphage
concentration decreased with distance from the source, but
only the coliforms showed a statistically significant decrease.
Fannin et al. compared the ratio of airborne coliphages to
coliphages found in sewage w'th the level of animal viruses in
sewage and calculated that, the expected concentration of
airborne animal viruses is 6.2-6.5 x 10~5/rn3 (Most Probable
Number).
In summary, the literature indicates that activated
sludge and trickling filter units emit potentially hazardous
bacterial aerosols and low levels of airborne animal viruses.
The concentrations of viable organisms generally decrease
rapidly with distance from the source, the rate of decrease
depending on such variables as wind speed, relative humidity,
solar radiation, air temperature, and obstructing vegetation.
267
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