WATER POLLUTION CONTROL RESEARCH SERIES
14010 FBZ 09/71
Concentrated Mine Drainage
isposal Into Sewage
Treatment Systems
:NVIRONMENTAL PROTECTION AGENCY • RESEARCH AND MONITORING
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WATER POLLUTION CONTROL RESEARCH SERIES
The Water Pollution Control Research Series
describes the results and progress in the
control and abatement of pollution in our
Nation's waters. They provide a central
source of information on the research,
development, and demonstration activities
in the Environmental Protection Agency,
through inhouse research and grants and
contracts with Federal, State, and local
agencies, research institutions, and
industrial organizations.
Inquiries pertaining to Water Pollution
Control Research Reports should be directed
to the Office of Research & Monitoring,
Environmental Protection Agency, Washington,
D. C. 20242
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Concentrated Mine 0 rain age Disposal
Into Sewage Treatment Systems
THE DISPOSAL OF ACID BRINES
FROM ACID MINE DRAINAGE IN
MUNICIPAL WASTEWATER TREATMENT
by
Environmental Research and Applications, Inc.
24 Danbury Road
Wilton, Connecticut 06897
for the
ENVIRONMENTAL PROTECTION AGENCY
Program #14010 FBZ
September, 1971
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EPA Review Notice
This report has been reviewed by the Office
of Research & Monitoring, EPA, and approved
for publication. Approval does not signify
that the contents necessarily reflect the
views and policies of the Environmental
Protection Agency, nor does mention of
trade names or commerical products
consitute endorsement or recommendation
for use.
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ABSTRACT
The effect of artificial iron-rich acid brines on municipal sewage
treatment processes was studied in small scale. The brines were
devised to simulate concentrates from treatment of acid mine drainage.
The processes included primary settling, activated sludge digestion,
and other unit processes.
The raw brines even at a level of 20% vol/vol or higher do not
interfere with primary settling, but activated sludge digestion is
completely inhibited by the acid. The brines when neutralized with
lime improve primary settling and the neutralized brines do not inhibit
activated sludge. Filtration of all sewage fractions is improved by
the lime-acid brine neutral mixture. At the very high concentrations
used, the neutralized brines give virtually complete removal of
phosphate from primary effluent, activated sludge effluent, or
anaerobic sludge digester decantate.
Cost of reverse osmosis membrane treatment of acid mine drainage to
produce the iron-rich acid brine is estimated to be in the range of
35C/1000 gal, of brine corresponding to original 48,000 gal. of acid
mine water or about 73C per thousand gallons of acid mine water
treated. Engineering analysis and costs are shown for transporting
the brine from the mine site to the sewage treatment plant by rail,
truck, and pipeline over distances ranging from 10 to 50 miles.
Transportation by pipeline involves per-mile costs on the order of
5C/1000 gal. of brine.
iii
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CONTENTS
Secti.OTi_
I Conclusions ]_
II Recommendations 3
III Introduction 5
Objective of This Study 5
IV Literature Review 7
AMD Treatment Processes 7
Phosphate Removal by Iron ^2
V Scope of This Study 19
Laboratory Investigation ^g
Engineering Analyses ^g
VI Methods 21
Artificial Acid Brine (AAB) 21
Experiments on Primary Settling 22
Experiments on Activated Sludge (AS) Digestion 23
COD-Digestion Experiments 23
Dosing Rate for AS Experiments 24
Anaerobic Sludge Digestion 25
Activated Sludge Digestion of Anaerobic Sludge 25
Digestion Decantate
VII Results & Discussion 27
Experiments on Primary Settling 27
Experiments on Activated Sludge Digestion 34
Experiments on Acid Shocking 36
COD Digestion by Activated Sludge 38
COD/BOD Relationship 38
Activated Sludge Digestion of Anaerobic Sludge 42
Digestion Decantate
Tertiary Treatment 42
Phosphate Removal 45
Biological Nitrification and Denitrification 47
Anaerobic Sludge Digestion 50
Anaerobic Sludge Digester Decantate 50
Digested Sludge Disposal 52
Sludge Conditioning 54
Oxidation of Ferrous Iron in AAB by Oxygen 54
Microbiological Status of AAB AS Cultures 57
Concentration of AMD by Reverse Osmosis 51
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CONTENTS (Con't)
Section Page
VIII Summary 69
IX Acknowledgements 71
X References 73
VI
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TABLES
Np_. Page
I Excess Sludge Production in Simultaneous and
Postprecipitation Units for Various Dosage
of Iron 13
II Composition of Artificial Acid Brine 21
III Recipe for Artificial Acid Brine 22
IV Summary of Settleable Solids Volume and Mixed Liquor
Total Suspended solids for Activated Sludge Cultures 24
V Composition of Stock Solution of SNP 24
VI Preliminary Experiment on AAB and Lime Effects on
Primary Settling 28
VII Further Preliminary Experiments on Primary Settling 29
VIII Effect of Ai-lD and Lime on Primary Settling 31
IX Effect of AAB and Lime on Primary Settling—
Decantate COB 32
X Effect of 10% AHD-I on Primary Settling The
pH Effect on the System 33
XI Effect of Concentration of MB on Primary Settling 35
XII Acid Shocking of Activated Sludge Cultures 37
XIII Effect of Neutralized AAB on Activated Sludge
Digestion 39
XIV COD:BOD Relationship of Various Waste-Treatment
Fractions 41
XV COD-Depletion Experiment on Anaerobic Sludge
Digestor Decantates 43
XVI Phosphate Removal from Primary Effluent by Lime
and AAB Treatment 45
XVII Fe, Ca, and Al Contents of Test Samples for
Phosphate Removal 46
vii
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TABLES (Con't)
XVIII Nitrogen Status of Sewage Fractions aac.
Wuhrmann
XIX Anaerobic Sludge Digestion in Digesters
Fed 100 ml Sludge Approximately Daily
XX Sunraary of Anaerobic Sludge Digestion Data
XXI Characteristics of Anaerobic Sludge Digester
Decantate
XXII Reduction of Dissolved Oxygen by Ferrous Iron
XXIII Change in Tine of Soluble Iron in AS containing
20% Neutralized MB
XXIV Oxygen and Work Requirements for Oxidation of
Ferrous Iron in Acid Brine
XXV Pipeline Transport Cost Estinates
48
51
52
53
5A
55
57
68
viii
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FIGURES
1 Flow Diagram of Complete Biochemical Oxidation
and Limestone. Neutralization Process 10
2 Solubility Diagram for Solid Phosphate Phase 16
3 Precipitation of Phosphate by Fe (III) 17
4 Activated Sludge Digestion of Anaerobic
Decantate 44
5 Disappearance of Soluble Iron in AS Containing
AAB Neutralized with Lime 56
6 Theoretical Oxygen Requirements For Oxidation
of Ferrous Iron 58
7 Flow Diagram for Concentrating of AMD - one MGD
Plant 62
8 Flow Diagram for Concentrating of AMD - 10 MGD
Plant 63
ix
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SECTION I
CONCLUSIONS
1. Many sites where acid mine drainage originates and degrades water
quality will require on-site treatment.
2. The products of treatment are concentrated acid brines or brine-
sludge mixtures and relatively pure water. Depending upon the method
of treatment, the product water can be suitable for safe discharge to
surface water courses or it can be a high grade water for industrial
use.
3. The brines produced by treatment are not of sufficient value to
pay for treatment by any known method, but they can be disposed of
through municipal wastewater treatment plants where they will be of
some value especially for sludge conditioning and nutrient removal.
A, Reverse osmosis offers an attractive method for on-site treatment.
5, It is economically feasible to transport acid brines from the
mine sites to municipal wastewater treatment plants by pipeline.
6. Some aspects of the development of an optimum scheme for disposal
of acid brines require further research especially nitrification,
denitrification, hydrolysis, and ultimate sludge disposal.
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SECTION II
RECOMMENDATIONS
1, Acid mine drainage sites should be inventoried or existing
inventories should be studied to find a site suitable for installation
of a pilot plant to produce acid brine (and pure water) for use at an
accessible municipal wastewater treatment plant,
2. A pilot scale investigation of the disposal of acid brine at a
municipal secondary wastewater treatment plant should be initiated.
The recommended project could use artificial acid brine until the
pilot plant brine becomes available,
3. Laboratory scale investigations of the use of acid brine for
certain advanced wastewater treatment unit processes should be
initiated. The processes should include biological nitrification and
denitrification, sludge conditioning, alkaline hydrolysis and
adsorption, and ozonation,
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SECTION III
INTRODUCTION
Acid mine drainage is a significant water pollution problem in
many mining areas of the United States, Acid mine waters originate
from the chemical oxidation of pyrites contained in coal and in shales
and other rocks associated with coal deposits and pyritic ore bodies.
In natural hydrographic regimes and especially in the regimes altered
by mining and other excavations, acid water find their way into water
courses. The acid and heavy metals in the drainage seriously damage
the stream flora and fauna even to the point of local extinction of
the biota. In the watercourse, acids are eventually neutralized
through reactions with natural alkaline materials and the heavy metals
are precipitated forming deposits of hydrous oxides of iron, aluminum
and manganese. In the course of this natural self-purification,
however, stretches of streams, miles in extent, can be degraded.
Objective of This Study
Various schemes for preventing the drainage of acid mine waters
into water courses have been proposed and undertaken, but not all
such problems can be economically solved by sealing the mine,
diverting the flow, or other such approaches (1, 2, 3), From the
standpoint of economic value, acid mine water are ordinarily too
dilute and too heterogeneous for product recovery (acid, iron,
other heavy metals) to be practical. Our study examined the feasibility
of concentrating acid mine water by certain processes and transporting
the concentrated brine to municipal waste treatment plants for disposal.
The relatively pure water produced from the treatment process would
be discharged to the local watercourse or used by a local municipality
or industry. With this broad objective—disposal of concentrated
acid brines via municipal wastewater treatment plants — we undertook
laboratory investigations of the effect of synthetic acid brine on
primary and secondary sewage treatment processes.
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SECTION IV
REVIEW OF LITERATURE
Introduction
The term acid mine drainage (AMD) is applied to drainages which
have an abnormally low pH value and to drainages which contain neutral
substances such as ferrous sulfate. Acid drainage from coal mining
operations usually contains sulfuric acid and ferrous, ferric,
aluminum and manganese salts in fairly high concentrations and also
calcium, magnesium and sodium salts, carbon dioxide and silicic acid.
Acid mine drainage is generated when the oxidation products of the
pyrite contaminate the surface drainage and the ground waters drained
from the mine. The feasibility of controlling the acidic contamination
of coal mine drainage at the source has been studied, discussed and
tried in the U.S.A. (4,5,6). Most of the suggested methods of control
have been based on the principle of excluding the mutual contact of the
three components of acid mine drainage: air, water, and pyrite. Some
of the segregation methods proposed were: diversion of water from acid
producing areas, exclusion of air from abandoned underground mines and
covering the pyritic materials with non-acid forming materials in
restored strip mine sites and refuse banks. It is evident from the
physical nature of the sources of acid contamination of coal mine
drainage that segregation will not be possible at some sites, and that
complete segregation will be difficult at many. It will be particularly
difficult to control the interaction of air, water, and pyrite in active
underground mine workings. Therefore the acid mine drainage has to be
treated after its formation and the cost of treatment is great. Thus
there is an increasing demand for cheaper mine drainage treatment
processes.
Acid Mine Drainage Treatment Processes
There are a number of processes reported in the literature for the
treatment of acid mine drainage and these are described in brief as
below:
1. Precipitation of Iron
By adding alkaline neutralizing agents such as hydrated lime,
sodium carbonate, and sodium hydroxide to acid solutions of
ferrous iron, iron can be precipitated. Other methods would
require expensive precipitating agents and are probably not
economical for high concentrations of iron.
2. Electrolysis of Iron (II) Solutions
Direct current hydrolysis of acidified iron (II) sulfate solutions
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produces a complex set of reactions and the three predominating
reactions are shown below:
H20 ------------- ^ 1/2 02 + 2H + + 2e~
Fe4"3 + 3H20 --- *• Fe (OH) 3 + 3H+3
The rate of removal of iron from solutions depends on the current
density at the electrode surface. Fe(OH>3 deposits as a brown
precipitate around the anode.
3. Aeration - Filtration
This is one of the most common methods used on a commercial scale
for removing iron from aqueous solutions. The process consists of
oxidation of iron (II) to iron (III) by aeration, followed by
hydrolysis of iron (III) and subsequent filtration
Fe +3 + 3H20 -------- K Fe (OH)3 (Fe203. n.H20) + 3H+
Superficially this process appears inviting for application to AMD
but many problems are involved. The two major ones are:
(i) conversion of Fe (II) to Fe (III) by air oxidation is slow
in media where the pH is less than seven.
(ii) lime or CaCO-j must be added to increase the pH. The con-
version of Fe (II) to Fe (III) by aeration is more rapid at
pH ~?>7 . However, for large volumes of AMD, neutralization
would require large amounts of CaCO-j.
4. Ultrasonic Methods
The use of ultrasonic energy in the oxidation of acidified iron
(II) sulfate has been investigated to a considerable extent with
some promising results. The rate of oxidation of iron (II) by
ultrasonic energy is a function of the intensity of the waves but
the relationship appears to be non-linear. Ultrasonic energy may
be useful in effecting coagulation of the hydrous iron (III) oxide
formed by oxidation.
5 . Ozone
Ozone treatment has been considered as a means of removing iron
from aqueous solutions. A combination ozone-activated carbon
system is presently in use for the removal of iron and manganese
from drinking water. (7)
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6. Irradiation and Photo-oxidation
Oxidation ef iron (II) using alpha, beta and gama radiation has
been studied and the mechanism is apparently the same as for
ultrasonic oxidation. The reactions appear to be about four
times faster in aerated water than in oxygen-free water. The
problems associated are: yield of the reactions are not partic-
ularly high and the depth of penetration of the water by the
irradiating particles is very short.
7. Lime Processes
The principle of the lime processes is that lime (calcium oxide
or calcium hydroxide) , or other strong alkali such as sodium
hydroxide , is mixed with the acid mine drainage to neutralize
the acids,
H2S04 + Ca(OH)2
and to precipitate the contaminating metal salts
Fe2(SOA)3 + 3 Ca(OH)2 ------------ + 3 CaS04 + 2 Fe(OH)3
The sludge formed by sedimentation from lime neutralization has
a high water content and presents a difficult disposal problem.
The cost of treatment by lime treatment is estimated at $1.09/
1000 U.S. gallons (1). Hanna et^ al (4) , in a review of the acid
mine drainage problem, concluded that neutralization processes
were not economically attractive except in specific cases
involving small well-defined areas.
8. Limestone Processes
This process used the minerals, limestone and dolomite (ie CaCO-j,
and Mg/CaC03) > and tne reaction with the acidic salt would be:
Fe2 (SOA)3 + 3 CaC03 + 3 H20 ------- > 3 CaSO^ + 2 Fe (OH)3+ 3 C02-
An upflow expanded bed process using limestone for the treatment
of acid wastes has been reported (8) . Sulfuric acid in concentra-
tions up to 5000 mg/1 could be neutralized without inactivation of
the limestone. The process was, however, stated to be unsuitable
for the neutralization of wastes containing iron salts because
of inactivation of the limestone.
There have been a number of cases (9,10,11) where the limestone
processes have been used. From a critical review of literature,
it is found that limestone has found application in the neutraliza-
tion of acid wastes other than those containing iron and other
precipitable salts, and that a considerable problem of limestone
inactivation must be expected when these salts are present.
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Air
Limestone
Grit
Acid Mine
Drainage*.
Flow
Balancing
-
Biochemical
Oxidation
-
Sedimenta*-
tion
— *
Limestone
Neutraliza-
tion
-*
Sedimenta- ]
tion
Active Sludge
ffluent
Filtration
M
09
ft
n>
Cake to Waste
Figure 1. Flow Diagram of Complete Biochemical Oxidation and Limestone
Neutralization Process
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"9. Biochemical Oxidation Processes
The simplest procedure to remove ferrous salts would be to
oxidize the ferrous salts to ferric which could be removed to-
gether with the original ferric salts by the limestone. Atmospheric
oxygen is by far the cheapest oxidant and is particularly attractive
since an excess of air could not produce too high a redox potential
in the treated drainage.
Unfortunately, atmospheric oxygen is not able to oxidize ferrous
salts in acidic solution despite the fact that the free energy
change is favorable (12). A micro organism capable of promoting
the atmospheric oxidation of ferrous salts in acid solution has
been widely reported in acid mine drainages (A, 13, 14). This
micro organism is an autotrophic bacterium, and Silverman and
Lundgren (15) reported the optimum growth temperature of this
bacterium to be 28°C, the optimum ferrous concentration to be
9000 mg/1 Fe, and the optimum pH range to be 2.5 to 4.0. The
bacterium studied by Silverman and Lundgren could not oxidize
manganous, nickelous, and cobaltous salts, but several authors
have reported the tolerance of acidophilic ferrous oxidizing
bacteria to metal ions such as manganese, copper, and zinc. (13)
A scheme proposed in England (4) for the control of acid mine
drainage pollution by biochemical oxidation and limestone
neutralization treatment is shown in Figure 1. The outstanding
characteristic of the complete process was the ease of control.
The required supervision was minimal since over-treatment in
the oxidation and neutralization stages could only be beneficial.
The practicability of the whole process to the treatment of acid
mine drainage has been demonstrated by this study. The limitations
of this process are: the dissolved iron content should be at
least 10 to 20 mg/1 and a total acidity of at least 25 mg/1 as
(CaCO-j) . The cost of treatment is estimated to range between
$0.67 to $1.90 per 1000 gallons.
10. Foam Fractionation
A foam fractionation technique (16) has been developed by Garrett
Research and Development Company, for removing phosphates and
suspended solids and lowering chemical oxygen demand. The
technique consists of treating the waste stream with ferric
chloride or alum, then dissolving air in the water under high
pressure (about 175 psi). When the pressure is released,
micro-size bubbles produce a creamlike foam that effectively
removes not only phosphates but also particulates in general
and both soluble and insoluble organic matter.
Tests with wastewater obtained at a Pomona, California, sewage
treatment plant showed that when alum or ferric chloride were
used in concentrations of 300 to 400 mg/1, foam fractionation
removed 90% of the phosphates and 80% of the suspended solids,
11
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and lowered the chemical oxygen demand by 70%. Preliminary
estimates for a 10 MGD foam - fractionation plant indicate a
capital cost of $235,000 and an operating cost of 6C/1000 gallons.
The land area requirements are very low, in the order of half
acre for a 10 MGD plant. As indicated in the preceding paragraphs,
none of these processes has yet been shown to be less expensive
for general application. In order to bring down the operating
costs of the sewage treatment facilities, attention has been
given to the use of waste materials, such as acid mine drainage.
Iron salts act as flocculating agents in the primary and secondary
clarification tanks. Iron salts act as an oxygen carrier in the
activated sludge process and the effectiveness of the returned
sludge is increased by the fact that sludge passing through the
cycle more than once becomes richer in iron, aluminum, and other
hydroxides and thus clarification of sewage and destruction of
protein are accelerated. In general it can be assumed that iron
may be helpful both as an oxygen carrier and as an adsorbent and
coagulant. (17)
The effect of iron salts on the removal of BOD, SS, phosphate are
discussed in the following sections:
Phosphate Removal by Iron
1. Elimination of Phosphate from Sewage
Phosphorus removal processes recently proposed are mostly based on
precipitation, with cations forming insoluble phosphate salts, or
on absorption by inorganic hydroxides. The reactions with Al3+
•31 l Or ' " 01 01
Fe° , and Ca or with combinations of FeJ and Ca^ offer some
economical possibility for the removal of phosphorus.
In addition to these inorganic chemical reactions, a biological
process might be considered theoretically, using the well-known
property of many microorganisms to store phosphates as poly-
phosphates in their cells when phosphorylation substances are
lacking. (18)
2. Iron (III) as precipitant
At pH>7, FePO^ results as a product, whose solubility product
is Ks = 10~ 23 at 25°C (18). Excess of Fe (III) is required in
the sewage treatment for the formation of a well-flocculating
hydroxide precipitate which includes the FePO^ particles. FePO^
particles act as an efficient absorbent for organic phosphate
compounds and eventually for polyphosphates.
Phosphate removal with Fe(III) has been considered exclusively
as an absorption process. This hypothesis contradicts the
opinion of Galal and Stumm, who favor a simple chemical precipi-
tation mechanism. (19) However, from a practical point of view,
12
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both reaction mechanisms lead to the same result and require the
same operating conditions.
Wben treating an activated sludge plant effluent, the stoichio-
metrical requirement 6f Fe is 1.8 mg/1 per mg/1 P. However, for
complete phosphate removal this has to be supplemented by at least
10 mg/1 for hydroxide formation. There are two methods of applying
the precipitants:
(1) Simultaneous precipitation
(2) Post precipitation
Simultaneous precipitation involves addition of the precipitant
directly to the influent of the aeration basin, promoting the
formation of hydroxide and binding of phosphates simultaneously.
Post precipitation involves the addition of precipitants to the
secondary effluents in a conventional manner. Both the methods
of applications have been studied on a laboratory scale (20) and
the following are the results:
A. Both processes required the addition of at least 20 mg/1
Fe3 for the reduction of phosphate from 7.15 mg/1 to 0.5 mg/1
or less.
B. Fe (OH)., flocculation in the simultaneous process was
frequently incomplete, resulting in an opalescent final effluent
with considerable carryover of colloidal Fe(OH)3 and phosphorus.
C. The addition of Fe (III) to the activated sludge resulted
in a striking change of the sludge biocenosis leading to better
settling.
D. When the dosing was more than 10 mg/1 Fe (III) , the protozoan
fauna disappeared completely within the first two days.
E. There was a remarkable increase of excess sludge volume in
the simultaneous precipitation units in comparison to conventional
treatment. (See Table 1)
TABLE 1
Excess Sludge Production in Simultaneous and Post Precipitation Units
(gallons per gallon of sewage treated) for Various Dosage of Iron (20)
Dosage of Fe3+ 10 mg/1 20 mg/1 30 mg/1
Simultaneous Precipitation 0.5 0.62 0.50
Post Precipitation 0.41 0.28 0.25
13
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According to the literature cited the hydroxide sludge formed is
of poor settling and dcwatering characteristics.
3. Removal of Phosphorus with Lime and Iron (III)
In order to reduce sludge solids production and excess sludge
volume in the precipitation unit, a combination of phosphate
precipitation with iron (III) and lime was tried. Iron was
added at 1.8 mg/1 Fe (III) for each mg/1 of phosphate and lime
added to raise the pH of the mixture to 8.8. This pH was
selected so that considerable precipitation of CaCO^ would take
place, which might be helpful as a thickening aid of the iron
hydroxide formed; at the same time the calcium carbonate was
thought to act as an absorbent for colloidal FePO^ and organic
phosphate. The results of an experiment run for 30 days with
the final effluent of an activated sludge plant demonstrated
that phosphorus removal was equivalent to the efficiency achieved
with lime or iron precipitation alone. Excess sludge quantity
and sludge volume were much smaller but the settling property of
sludge was poor. Essential factors to be considered for the
operation of this process are: the amount of orthophosphate to
be removed, the alkalinity, and the Ca hardness of the treated
sewage. This combined process is expected to be the most economic
over a wide range of conditions. The quantity of precipitation
chemicals are moderate, and the sludge produced thickens rapidly
and can be effectively dewatered to a highly concentrated slurry
or a filter cake, which may be dumped without danger of secondary
leaking of phosphorus.
Alum and Iron (III) precipitation have two disadvantages: the
costs for the chemicals are relatively high, and the processes
lead to a difficult sludge disposal problem. In the case of alum
precipitation, the recovery of the precipitant as proposed by Lea
et al (II) , may lighten the cost burden to some degree.
(a) Phosphate Precipitation
+2 +3
The affinity between multivalent aqueous metal ions (Ca , Fe t
Al+3) and PO^ ~3, HPO^"2 or polyphosphate is about 4 - 12K cal per
mole. With simple solutions of orthophosphate, well defined
reaction products such as FePO^. 21^0 (Strengite) can be formed
in accordance with the stoichiometry of the reaction. Organic
phosphorus compounds are surface-active and may become adsorbed
on precipitates and other interfaces, but counter-ion adsorption
alone cannot account for the high removal efficiency because of the
stoichiometry observed. A substantial fraction of the phosphorus
in sewage is contained in the form of suspended matter; this fraction
is removed by coagulation rather than by precipitation. Ferric
salts are good coagulants and can be used to coagulate the phosphate
colloids.
14
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(b) Solubility Relations
The continuous precipitation of phosphate in treatment of wastes
is a dynamic process; most likely, equilibrium conditions are not
attained. The formulation of the stoichiometry of the chemical
reaction defines the minimum requisite quantity of chemicals. The
equilibria (21) that need to be considered are listed below:
1. Solubility equilibra
>•
(log K 25°C) = -23)
2H20 (s) ------ >• Fe+3 + PQ-3 +
2. Acid - Base Equilibria
Fe+3 + OH- ------------ fr-FeOH+2
FeOH+2 + OH~ ----------- -
Fe(OH)+ + QH~ --------- >>Fe(OH)3 (s)
3. Complex Formation Equilibria
FeHPO,+ -------------- *-Fe+3 + HPO,~2
4 4
FeH2P04+2 ------------- >Fe+3 + H2PO~
Fe (HP207)2~3 -------- ^Fe+3 -1- 2HP 0?~2
(c) Kinetics of Phosphate Precipitation
Equilibrium calculations and experimental verification show, that
FePO^ (s) is a stable solid phase if phosphate is precipitated
in the low pH range of 4 to 6 (See Figure 2). Precipitation of
phosphate with Fe (III) is very fast. The rate of removal is
controlled by agglomeration of colloids and by settling.
Precipitation of phosphate by Fe (III) will be more efficient if
the Fe (III) is produced by slow, homogeneous generation of ferric
ion through the oxidation of ferrous ion, than it will be if Fe
(III) as such is added to the wastewater. In the latter case, ferric
ions, because of relatively high localized concentrations, tend to
react (depending on pH) with OH~ ions rather than with phosphate
species. If, on the other hand, ferric ion is produced homogeneously,
each ferric ion, as soon as it is being formed, comes into contact
and reacts with phosphate species. The better scavenging effect
of homogeneously formed Fe (III) when compared to Fe (III) introduced
directly is shown in Figure 3.
15
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0)
u
fl
to
o
JS
a.
M 8
o
10
FeP04
PH
10
12
Figure 2. Solubility Diagram For Solid Phosphate Phase
16
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10-.
With Fe (III)
Salt
Fe (III) Homogeneously
Generated
Log
-4
Fe (III) Added
Figure 3. Precipitation of Phosphate By Fe (III)
17
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SECTION V
SCOPE OF THIS STUDY,
The subject study consisted of a literature review, laboratory
experiments, and engineering analyses. The general literature
review is given separately above and other previous published work
is cited at appropriate places throughout this report.
Laboratory Investigation
The wastewater treatment processes studied were primary settling,
activated sludge digestion, and anaerobic sludge digestion. The
general approach was to compare the activity or effectiveness of
acid-brine treated laboratory units with control (untreated) units
with respect to COD removal, suspended solids removal, and phosphate
removal. The broad question addressed was whether the unit process
was degraded or enhanced by acid brine and neutralized (with lime)
acid-brine.
Engineering Analyses
The general thrust of engineering analyses was economic. What
are the estimated costs for brine production at the mine site; what
are the costs of transporting brine via pipeline or by tankbarge,
tanktruck, or railway tankcar; what is the value (if any) of acid-
brine at a municipal wastewater treatment plant; what are the costs
for lime or other basic salts for neutralizing acid-brine?
19
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SECTION VI
METHODS
For analyses of pH, acidity, alkalinity, iron, dissolved
oxygen, BOD, COD, etc., Standard Methods (22) or simple modifications
of standard methods were used. Raw sewage, primary sludge, digestor
sludge, and primary effluent were obtained from the City of Norwich,
Conn, sewage treatment plant, The Norwich sewer system is virtually
completely combined stormwater and sanitary wastewater with some
industrial wastewater including those from textile dying and finishing
and food manufacturing. There is considerable day-to-day variation
in the raw wastewater and in the quality of the primary effluent. The
sewage is ordinarily faintly acid pH 5.5-6.5 with a total alkalinity
of about 100 mg/L (as CaC03>.
Artificial Acid Brine
The recipe for artificial acid brine (AAB) was devised to mimic
a typical natural acid mine water concentrated about fifty-fold by
a membrane process or any other suitable process (see Tables II, &
III).
In order to simulate more ideally the composition of acid
brines produced from natural acid mine drainage, all salts used should
be the sulfates, because acid mine water is very low in other anions
including chloride. The use of the magnesium and ferric chlorides,
however, does not obviate the validity of any results obtained and
reported here. The processes investigated are not sensitive anions
over the range used in the present work,
Table II
Composition of Artificial Acid Brine
pH 2.0
acidity 5000 mg/L as CaC03
ferrous iron 1000 mg/L
ferric iron 1000 mg/L
magnesium 250 mg/L
aluminum 250 mg/L
manganese 100 mg/L
calcium sulfate saturated
21
-------
The receipe used to achieve that composition is given in Table 3.
The stock solution is labeled AAB-I.
Table III
Recipe for Artificial Acid Brine
Ferrous sulfate FeSO^-7H20 5,0 gm
Manganous sulfate MnSO,-H^O 0.3 gm
*r •&
Potassium alum A1K(SO^)2'12H_0 4.4 gm
Ferric chloride FeCl3«6H20 4.8 gm
Magnesium chloride TIgCl «6H 0 2.1 gm
Make up to 1.0 liter with water saturated with calcium sulfate and
adjust to pH 2.0 with sulfuric acid. The solubility of calcium
sulfate dihydrate at room temperature is on the order of 3 grams per
liter; a saturated solution has a pH of 6.3. Adding sulfuric acid
to make the solution pH 2.0 does not increase the solubility because
the sulfate added decreases the dissociation of the salt. The recipe
above with sulfuric acid added to bring the pll to 2.0 has a titrated
acidity of 7800 mg/L (as CaCO-j) .
A second recipe (AAB-II) was used with ferrous iron (as the
sulfate) added to make the total iron 10,000 mg/L. The second AAB
formulation had a titration acidity of 23,000 mg/L (as CaC03). AAB
stock solutions were stored under ambient conditions and no appreciable
oxidation of ferrous iron occurs from contact with air because of the
low pH. (2.0).
Experiments on Primary Settling
Using Standard Methods, total suspended matter was determined on
various sewage treatment fractions by settling in an Imhoff cone and
by membrane (0.45 microns porosity) filtration. For some purposes
gravity filtration through coarse paper was used to obtain filtrates,
a procedure which removes somewhat more suspended matter than primary
settling. For obtaining decantate from laboratory anaerobic sludge
digesters, centrifugation at 1800 rpm for five minutes was used. That
speed corresponds to a relative acceleration value of about 600g. The
centrifugation procedure produces a decant about equal in quality
to gravity filtration through coarse paper. We have noted the type of
separation procedure used in specific instances in the Discussion section.
22
-------
Experiments on Activated Sludge Digestion
Laboratory units for activated sludge digestion (AS) were
polyethlene battery jars of twenty-liters capacity containing about
ten liters of mixed liquor. AS cultures were stirred with a one-
inch disc impeller at about 1000 rpm and were aerated with a single
tube with a capillary tip at the bottom of the culture. The aerators
were set to deliver about 750 ml/min of air across a column depth of
nineteen centimeters. The surface area of the cultures is 150 cm2.
These parameters combine to give an aeration rate of 50 L/m2.min or
0.17 ft3/ft*.min. Dissolved oxygen measurements with a Yellow Springs
Instrument Co. Model 54 dissolved oxygen meter indicated that dissolved
oxygen in the AS cultures would fall to 40-60% of saturation following
batch feeding and would rise to virtual saturation within one hour.
AS cultures were not thermostated; ambient temperatures varied
between 60 and 85°F (16-29.5°C).
Suspended (settleable) solids levels in control AS cultures were
maintained at about 40 ml/L, and in the AAB-fed cultures, at 190 ml/L,
The average ratio of settleable solids volume by Imhoff cone (ml/L)
to total suspended solids by membrane filter (mg dry/L) was 12.3 for
control cultures and 10.8 for AAB-fed cultures. For 53 samples of
control AS mixed liquor, the average total suspended solids was
492.5 mg dry/L and for seventeen samples of AAB-fed cultures the
average was 2094.0 mg dry/L. -These data are summarized in Table IV.
The ratio of total suspended solids to settleable solids is not a
valid measure of the dry weight per unit settled volume of the sludge
because the total suspended solids also contains some so-called residual
suspended solids or non-settleable suspended solids. Typical values
of residual suspended solids—the total suspended solids by membrane
filter of the Imhoff cone decantate—are 70 mg dry/L for control AS
cultures and 190 mg dry/L for AAB-fed.
COD-Digestion Experiments
In a typical COD-digestion experiment, one or more control AS
cultures would be fed a dose of primary effluent and the AAB-AS
culture would be fed the same dose with twenty percent by volume
of AAB. Samples of mixed liquor would be withdrawn periodically
and settled in Imhoff cones, and the COD of the decantates would be
determined. This was done day after day. A single experiment would
involve taking the decantate COD before feeding, the decantate COD
immediately after feeding, called the zero-time COD, and the decantate
COD after two, four, six hours and 20-24 hours—the following day.
In many experiments, only the 20-24 hour decantate COD was taken
because of the difficulty in handling many COD analyses in a few hours
time. In our experience, COD-digestion is virtually complete in
four to ten hours. Corresponding BOD^'s were also taken in some
instances. The BOD/COD relationship is discussed below. It was
23
-------
promptly learned that raw AAB completely inhibited AS digestion, and
from that time on all COD digestion experiments were done with AAB
neutralized with lime (Pfizer-Nelco High Calcium lime containing
73.0% CaO and 0.5% MgO).
Table IV
Summary of Settleable Solids Volume and Mixed Liquor
Total Suspended Solids for Activated Sludge Cultures
53 samples 17 samples
Control Cultures AAB-fed Cultures
Settleable Total Suspended Settleable Total Suspended
Solids Solids Solids Solids
(ml/L) (mg dry/L) (ml/L) (mg dry/L)
Range Ave. Range Ave, Range Ave. Range Ave.
20-70 40.1 200-900 492.5 150-260 193 1500-2400 2094.0
492.5/40.1 = 12.3 mg dry/ml 2094.0/193 = 10.85 mg dry/ml
Correcting for non-settleable solids (see text)
492.5-60/40.1 = 10.8 mg dry /ml,' 2094-190/193 = 9.8 mg dry/ml
J)osing Rate for AS Experiments
As has been mentioned, the Norwich sewer system is combined storm
and sanitary. Thus during periods of heavy rains, the sewage becomes
very dilute and at those times the primary effluent used as feed for
AS cultures was augmented with a synthetic mixture of sucrose, potassium
phosphate, and ammonium carbonate formulated to give a ratio of C:N:P
of approximately 7:1:0.05 or 140:20:1. The recipe for SNP is given
in Table V.
Table V
Composition of Stock Solution of SNP
Sucrose 250 gm/L
(NH4)2 C03.H,0 50 gm/L
KH2P04 3.0 gm/L
SNP solution has a COD of about 250 mg/ml.
24
-------
In a typical COD-digestion experiment four liters of mixed liquor
would be drawn off and settled for sludge recovery and return (if
necessary to maintain the design level), and for determination of the
decantate COD before feeding. A four-liter batch of feedstock was
prepared by mixing water, primary effluent of known COD, and SNP
(if necessary) to bring the zero-time (after feeding) COD to about
400 mg/L. A typical calculation for the feed mixture is shown below.
CODzero - t ^ 400 - 400 = 4 x C°Dfeed + C°Dbefore
10
before was typically about 60 for control AS and 120 for
AAB-fed. Thus for a nominal value of 100, feed would be 850 mg/L.
If primary effluent were higher than that, it would be diluted with
water; if lower, SNP would be added,
Anaerobic Sludge Digestion
Two anaerobic sludge digesters for four-liters capacity were
charged with 400 ml heavy primary sludge plus 2500 ml water. Thereafter
one was maintained on control activated sludge and the other on AAB-fed
activated sludge. The digesters were kept closed and stirred with a
magnetic stirring bar. The digesters were not heated; the ambient
temperature varied between 65 and 75°F (18.5 and 24°C). The head-space
in the digestors was 900 ml. Every morning the anaerobic cultures
were removed from the stirring motor and permitted to settle. A 100-ml
portion of decant was removed for analyses and 100 ml of settled
activated sludge was added to restore the original volume. Decant
samples were analyzed for pH, iron, COD, occasionally BOD, and total
phosphorus. Occasionally small samples of mixed liquor were withdrawn
for total suspended solids determinations to follow the course of
sludge accumulation and/or digestion. Toward the end of the period of
experiment, it was deemed expedient to obtain decantate by centrifugation.
Activated Sludge Digestion of Anaerobic Sludge Digestion Decantate
Conventionally, digested sludge is either drawn off and dried on
sand beds for land-fill disposal or it is dewatered on a vacuum drum
filter.!. In the latter case the filtrate, which is very high in BOD
and COD is recycled through the plant. In order to determine the
treatability of anaerobic sludge digestion decantate, we centrifuged
the contents of our two digestors. The decantate from the control
digestor had a COD of more than 4000; the decant from the sludge
digester fed iron-rich activated sludge had a very low COD—on the
order of 100. The former (control) decantate was passed through the
AAB-fed AS digestor without added AAB, and the latter (iron-rich)
decantate was passed through a control AS digestor augmented with SNP
synthetic feed-stock to a COD level equal to our conventional batch
25
-------
feed. The purpose of the latter test was merely to determine whether
the iron-rich anaerobic decantate was toxic to AS. As discussed
latfer, the decantate proved to be nontoxic.
26
-------
SECTION VII
RESULTS AND DISCUSSION
Experiments on Primary Settling
Experiments on sludge settling were carried out in Imhoff cones.
The results of a preliminary experiment on the effects of AAB, lime,
and AAB plus lime are shown in Table VI.
On the occasion of the preliminary experiment raw Norwich sewage
contained only a small amount of settleable solids (this is frequently
the case). Lime alone improves settleability markedly, if sufficient
lime is added to bring the pH to 10.5. That effect may be due in
part to precipitation of sesquioxides. It is not shown in the Table
(VI) but the quantity of acid required to neutralize the lime treated
sewage from pH 10.5 to pH 8.4 (as if for stream discharge from primary
sewage treatment only) is equivalent to 50 mg/1 (as CaCO^). Coincident
tally, that is equal to 50 mg H^SO^/L. AAB alone enhances solids
settling appreciably, albeit this response is much more variable than
with lime alone. AAB alone causes some coagulation of suspended matter
but the clots are bulky and sludge volume shown in Table VI is some-
what misleading. Neutralizing the AAB-sewage mixture is equivalent
to adding AAB plus lime, in which case a hydrous iron oxide floe forms,
which scavenges colloid and other particles as it settles. The
addition of AAB plus lime produces a copious sludge, which settles very
rapidly (a few minutes), and produces a clear effluent, albeit a pale
yellow-colored one.
Another preliminary experiment is shown in Table VII.
The data of Table VII confirm in a general way the results of
earlier preliminary experiments. The improvement in clarity of the
supernatant from AAB plus lime is especially noteworthy. The addition
of AAB at a level higher than 10% AAB-I vol/vol as shown in sample #6
does not improve decantate clarity; AAB-II at 5% vol/vol is equivalent
in its iron content to 25% AAB-I. In these experiments turbidity was
measured with a Klett-Summerson industrial model colorimeter using a
#42 blue filter. The decantate from AAB/Lime treated samples is very
clear but it does have some yellow color; except for highly colored
solutions, however, the error is not significant. In order to check
the sludge volume due to iron precipitates themselves and to determine
the apparent turbidity of the decantate due to AAB itself, a sample of
AAB 10% vol/vol in water was neutralized with lime to pH 7.6 and the
sample was settled in an Imhoff cone in the usual way. That sample
is shown as sample #7 in Table VII. The sludge volume of 250 ml, higher
than that of sample #5 (180 ml), indicates either that precipitation
of iron hydrous oxide is inhibited somewhat by substances in sewage or
that the solids in AAB/lime treated sewage settle more densely than
27
-------
NJ
co
TABLE VI
Preliminary Experiment on AAB and Lime Effects on Primary Settling
mple It
1
2
2A
3
3A
4
5
6
6A
7
Treatment
none
none
none
Lime
Lime
(200 mg/L)
Lime
AAB-I + Lime
(10% vol/vol)*
AAB-I
(10% vol/vol)
AAB-I
(10% vol/vol)
Lime
(250 mg/L)
PH
5.7
5.7
6.0
9.0
9.7
7.0
7.1
3.0
3.0
10.5
Volume of Sludge
Settled (ml)
4.0
5.0
2.0
12.0
6.0
6.0
80.0
40.0
12.0
30.0
Clarity of
Supernatant
poor
poor
poor
poor
fair
poor
good
fair
fair
fair
* AAB-I + Lime (10% vol/vol) means the part lime neutralized AAB-I to nine parts sewage,
-------
TABLE VII
Further Preliminary Experiments on
Primary Settling
Sample //
Treatment
PH
Volume of Sludge
Settled (ml)
Relative Turbidity
of Supernatant
VO
1
2
3
4
5
none
Lime
(100 mg/L)
MB -I
(10% vol/vol)
AAB-II
(5% vol/vol)
AAB-I + Lime
(10% vol/vol, lime
6.8
7.5
2.4
2.8
6.0
12
15
23
32
180
to pH 6.0)
AAB-II + Lime
(5% vol/vol, lime 6.0
to pH 6.0)
AAB-I + Lime
(10% vol/vol in water) 7.6
190
250
225
200
190
225
35
45
-------
the hydrous iron oxide in lime-neutralized AAB. Judging from the
appearance of the settled hydrous iron oxide in the latter case, we
prefer the interpretation that the sewage-solids/hydrous-iron-oxide
mixed solids are denser. The turbidity of the decantate from sample
#7 was 8 units; thus, the turbidity from suspended solids of samples
#5 and #6 can be considered to be slightly lower than the measured
turbidities shown in Table VII. For comparative purposes, a kaolin
suspension prepared according to Standard Methods gives a Klett
turbidity of 4.5 units per mg kaolin/L.
The results of a third preliminary experiment on primary settling
are shown in Table VIII.
The results confirm earlier experiments. Especially noteworthy are
samples #5 and #6 which show that amounts of lime-neutralized AAB in
excess of 10% AAB-I do not give any further improved clarification of
sewage, and also sample #4, which shows that lime in a quantity of
sufficient to raise the pH of the Norwich sewage to pH 9.0 significantly
improves primary settling. In other trials with lime alone, we observed
that settled sludge volume was approximately doubled by lime treatment
to pH 10 or slightly higher.
The results of a definitive experiment, in which COD analyses of
primary settling decantate were carried out, are given in Table IX.
Unlike earlier experiments, this experiment showed good clarification
from the highest level of lime-neutralized AAB tested (50% AAB-I
equivalent to a total iron concentration of 1000 mg/1 in the AAB/sewage
mixture). Of course, something less than half of the clarification
is due to the simple dilution effect. It should be noted that the
turbidity units are given in terms of the kaolin equivalent in mg/1;
for comparison with earlier reported turbidities in Klett optical
density units multiply by 4.5. Thus the sewage used in this experiment
had a Klett turbidity of 330 units in comparison to 160 and 225
units in earlier experiments. On the other hand the sewage had
virtually no settleable solids as shown by sample #1.
The COD removal by 10% neutralized AAB-I (sample #4) was 23%, and
for 50% neutralized AAB-I (sample #6), nil. At the lower concentration,
neutralized acid brine can remove some suspended and soluble COD not
removed by ordinary primary settling.
In order to determine the effect of pH on the decantate clarity
and COD-removal in AAB-treated sewage, an experiment using 10% vol/
vol AAB-I was run. The results of the experiment are given in Table X.
From the data it is clear that the optimum lies between 6 and 8.
That conclusion agrees with Packham's (23) results on water clarifica-
tion using alum. Packham determined the amount of alum needed to remove
50% of various colloidal suspensions (kaolin, other clays, calcite, and
quartz) as a function of pH. The minimum quantity of alum for 50% removal
30
-------
TABLE VIII
Effect of AMD and Lime on Primary Settling
Sample # Treatment pH Sludge Volume Turbidity of
(ml) Decantate
8 none 6.0 7.5 160
1 AMD-I
(10% vol/vol) 2.8 12 90
2 AMD-I
(50% vol/vol) 2.0 1 135
3 AMD-II
(10% vol/vol) 2.6 7.5 160
4 Lime 9.0 8 90
5 AMD-I + Lime
(10% vol/vol; Lime to pH 7) 7.0 200 65
6 AMD-I + Lime
(50% vol/vol; Lime to pH 7) 7.0 350 95
7 AMD-II -f Lime
(10% vol/vol; Lime to pH 7) 7,0 400 175
-------
TABLE IX
Effect of AAB and Lime on Primary
Settling—Decantate COD
ro
mpl
1
3
2
4
5
6
e # Treatment pH
none 7 . 1
Lime alone 10.0
(250 mg/L)
AAB-I alone 3.0
(10% vol/vol)
AAB-I + Lime 7.0
(10% vol/vol; 1100 rag Lime/L)
AAB-I alone 2.4
(50% vol/vol)
AAB-I + Lime 6.8
Settled Sludge
Volume (ml)
< 1.0
10
12
230
<1.0
400
COD*
mg/L
945
854
900
735
1000
1000
Turbi<
(Equfr
Suspei
73
63
105
13
86
21
(50% vol/vol; 5650 mg Lime/L)
Corrected for dilution and COD of ferrous iron added
-------
TABLE X
Effect of 10% AMD-I on Primary Settling &
The pH Effect on the System
pH of decant
Lime requirement
(mg/L)
Settled volume
£ (ml)
Turbidity (mg/L)
Decantate COD
(mg/L)
1
7.0
0
0
'5
5
2
2.9
0
24
100
644
3
4.3
310
120
45
550
Experiment
4
5,6
610
172
13
530
5
6.2
1000
172
17
494
6
8.3
1250
230
8
494
COD removal (%)
25
35
40
45
45
-------
of all the colloids fell between 6.5 and 7.5. Morgan and Stumm (24)
suggested that hydrous iron oxide suspended particles bind anions
and colloids optimally at a pH range of 5.0 to 8.0.
To determine more precisely the effect of concentration of
neutralized AAB on primary settling, an experiment was conducted
using AAB-I at 5, 10, and 25% levels, and AAB-II at 5 and 10%. To
obviate the necessity for making corrections on COD analyses, COD
measurements were made on the gross mixtures (before settling) and
on the decantates after settling. The results of the experiment are
given in Table XI.
On the assumption that acid brines would only be disposed of in
municipal wastewater treatment plants with secondary or higher grade
treatment, the turbidity data is only of interest as a very rough
measure of removal of suspended solids. Based on corrected turbidity
measurements, the optimum level of AAB is AAB-I 10%; the corresponding
level of AAB-II is 2%. In terms of COD removal, the highest AAB
levels tested gave the best results. It should be stressed, however,
that higher levels of AAB require larger amounts of lime for
neutralization. Furthermore, the sludge volume to be disposed of,
mostly hydrous iron oxides, is considerably greater at the higher
AAB levels. For example, for 25% AAB-I or 50% AAB-I, the settled
sludge volumes are 50% and 45% of the volume of raw sewage treated,
respectively.
The general conclusions drawn from the above experiments are that
primary settling is improved by the addition of AAB at a level as
high as 10% vol/vol neutralized with lime to a pH in the range of 6.0
to 8.0, that the lower end of the pH range calls for less lime, and
that primary settling is not impaired by AAB in the range 10-15%
even without neutralization. As will be shown later in this report,
neutralization of acid in added AAB is necessary for proper functioning
of activated sludge. The option of neutralizing before or after
primary settling is however open to the operator. If primary sludge
is combined with settled activated sludge for disposal, however, the
option is academic.
Experiments on Activated Sludge Digestion
Activated sludge cultures (AS) were operated on a daily batch
basis on weekdays during October, November, and December. A series of
experiments on acid (AAB) shocking of AS cultures was carried out to
determine the safe interval between adding AAB and neutralizing to the
physiological range of pH. It should be noted that the alkalinity of
the primary effluents we used was only about 100 mg/1 as CaCOn which
is equivalent to only 13 ml of AAB-I stock solution. If more than that
quantity of AAB is added to primary effluent, the pH drops below the
physiological range. Other experiments were carried out to compare
AS digestion in control cultures with that of a culture fed lime-
neutralized AAB. A third series of experiments was done to determine
the AS digestibility of decantate from anaerobic sludge digesters.
34
-------
TABLE XI
Effect of Concentration of AAB on Primary Settling
U)
Ui
mple
// Treatment pH Lime Turbidity Kaolin
Required Equivalent
(rag/L) (mg/L)
COD Settled
Before Decantate % Sludge
Settling (mg/L) Removal Vol. (ml)
(mg/L)
Uncorr. Corr,
for Color
1
2
3
4
5
6
None 6.5 90 90
AAB-I 7.2 300 325 44
(5% vol/vol)
AAB-I 7.2 1000 89 27
(10% vol/vol)
AAB-I 7.6 2000 74 59
(25% vol/vol)
AAB-II 7.2 1000 205 163
(5% vol/vol)
AAB-II 7.2 2500 280 214
1550 1440 7 4
1550 1135 27 40
1550 1025 34 170
1760 950 54 380
1480 1010 32 280
1580 920 42 400
(10% vol/vol)
-------
Experiments on Acid Shocking
The results of acid shocking experiments are given in Table XII.
The experiment shown was done at low loading; but it is obvious that
exposure to acid for only one-half hour does not seriously impair
digestion. Exposures for one hour or longer does inhibit digestion.
In a second experiment a culture was exposed to acid for 24 hours
and then neutralized and fed a portion of strong waste sufficient
to bring the initial COD to 1000 mg/1. Both the control and the
acid-exposed AS culture gave a 24-hour^ COD of about 250 mg/1 or
about 25% COD removal in 24 hours. A third experiment at low loading
was carried out on a culture exposed to acid for 24 hours (pH 3.2).
After the exposure, the culture was neutralized and fed along with a
control. The results are given below. Although digestion is obviously
slowed down for 16 hours, after 24 hours a comparable amount of digestion
is obtained.
Control Acid-Exposed
COD % removal COD % removal
initial 185 148
15 hrs. 73 60 129 13
18 hrs. 92 111
24 hrs. 75 60 75 50
In a fourth experiment, COD digestion was followed for the first
two hours after neutralizing and feeding an AS culture exposed to acid
for 24 hours. The results are given below:
Control Acid-Exposed
COD % removal COD % removal
initial 310 292
20 min. 256 292
60 min. 220 256
120 min.183 40 256 12
Thus, COD digestion begins within an hour of neutralizing and feeding
but the inhibition is quite marked.
The general conclusions to be drawn from the above experiments are
that AS cultures can be exposed to AAB at concentrations sufficient to
reduce pH to below 3 for one-half hour without significantly affecting
36
-------
TABLE XII
Acid Shocking of Activated Sludge Cultures
Exposed to AAB Control
Time exposed before neutralizing
u>
initial
2 hr.
4 hr.
6 hr.
24 hr.
1/2 hour
115
73
75
36
__
1 hour
115
75
73
85
67
4 hours
COD
115
91
85
—
75
133
49
36
29
__
-------
the COD digestion capability. Longer exposures damage the digestion
capability, but activity recovers after neutralization. The time
required for recovery depends upon the time of exposure and on the
loading, which introduces new active microbes along with digestible
organic matter. Even with an exposure to acid of 24 hours and with a
low loading following neutralization, recovery is virtually complete
in 24 hours. Even before recovery, a reasonably clear effluent with
some COD removed can be produced in the damaged culture because of
the improvement in suspended solids removal due to hydrous iron oxide.
COD Digestion by Activated Sludge
A comparison between the day-to-day performance of control AS
cultures and a culture fed neutralized AAB is given in Table XIII.
Table XIII shows the initial COD's, the COD after 24 hours aeration,
and the % removal of COD in the two types of AS cultures. Some
typical COD-depletion curves are given in a later section (COD digestion
by AS on anaerobic sludge digestion decantate), but for present purposes
it is sufficient to say that 85-90% of the 24-hour digestion occurs in
the first ten hours. From the data it is clear that the performance of
the AAB-fed culture is about the same as that of the control in terms
of daily COD removal. The lower COD removal (average 64.3%) in the
AAB-fed culture in comparison to the control (average 70.8%) may be due
in part to the fact that the AAB-fed culture was loaded in the average
at a rate about 6% lower than the controls, without regard to the
solids levels in the cultures. The settleable solids level in the AAB-
fed culture was maintained at about 190 ml/L as compared to about 40 ml/L
in the controls, but the greatest part of the solids in the AAB-fed
culture was hydrous iron oxide (and other hydrous sesquioxides). Judging
from mixed-liquor COD and BOD analyses the organic solids contents of
both types of AS cultures were about equal, at around 500 mg dry solids/L.
We note a rough trend in the data on COD removal in which the
performance in the second half of the period is poorer than in the first
half while just the opposite seems to be the case in controls. In the
latter case, one can attribute the trend to adaptation to the substrate;
in the former case we would attribute it to an increased turbidity in
the one-hour decantates. The turbidity seemed to be due to very small
coccoid bacteria with a thin rind of iron oxide, but we would not press
that point too far. With longer settling times, e.g. four hours, the
AAB-fed AS cultures produces a remarkably clear decantate. Another
important characteristic of the AAB-fed activated sludge is its filter-
ability. Ordinary activated sludge clogs a laboratory vacuum filter
(paper or membrane) in a matter of seconds, whereas the AAB-fed AS
filters rapidly through a thick pad of sludge. The same can be said of
the anaerobically digested sludges.
COD/ uiQD Rela tion ship
In order to translate treatment effectiveness measured by COD
38
-------
TABLE XIII
Effect of Neutralized AAB on Activated
Sludge Digestion
Control AAB
Oct.
Oct.
Oct.
Oct.
Oct.
Nov.
Nov.
Nov.
Nov.
Nov.
Nov.
Nov.
Nov.
Nov .
Nov.
Nov.
Dec.
Dec.
Dec.
Dec.
13
16
19
28
30
2
4
5
6
10
12
13
17
18
19
23
9
14
16
21
COD
o
475
380
340
228
446
273
358
415
334
242
309
405
486
300
209
300
366
1265
495
COD24 ACOD % COD COD24
removal
110
190
130
91
102
86
68
135
140
46
101
50
50
56
37
38
51
—
84
Average
365
190
210
137
344
127
290
280
194
196
208
355
436
244
172
262
222
774
329
39
77
50
62
60
77
46
83
68
58
81
67
88
89
81
82
87
61
61
67
To
335
226
170
177
464
213
307
364
296
284
329
373
465
275
190
280
315
925
401
, 357
78
170
100
55
18
102
137
51
51
70
120
134
120
175
93
113
91
144
124
133
ACOD %
removal
165
126
115
159
362
76
256
313
226
164
195
253
290
182
77
189
171
521
277
224
49
56
68
90
78
36
83
86
76
58
59
67
62
66
41
68
54
56
69
"6*473
-------
changes into BOD terms, both waste-strength parameters were measured
on various waste treatment fractions—primary effluent, AS mixed liquor,
AS decantate, and anaerobic sludge digestor decantate. The results of
the analyses are given in Table XIV. As can be seen in the data, the
Norwich primary effluent has a lower COD than BOD, the average COD:BOD
ratio being 0.72. In oxidation of organic matter by acid chromate,
the oxidation proceeds only to the stage of low molecular weight organic
acids; in other words, low molecular weight organic acids are refractory
in the COD analysis. The Norwich sewage is almost always slightly acid
and has a sour or ester-like odor. Control AS mixed liquor shows an
average COD:BOD ratio of 1.46, which is typical of a general mixture of
biological materials. Cell walls and cellulosic matter are nearly
completely oxidized by acid chromate, but only incompletely in a 5-day
BOD incubation. The CODrBOD ratio is higher in the AAB-fed mixed
liquor (1.93) indicating more biorefractory materials in the sludge.
The decantates from AS cultures have even higher CODrBOD ratios, a
reflection of the fact that biological sorption and oxidation of sub-
strate has taken place in the AS cultures. Decantate from the anaerobic
sludge digestor fed control sludge has a COD:BOD ratio of about 1 and
even more noteworthy a very high COD and BOD. The anaerobic sludge
digestor fed iron-rich sludge (from the AAB-fed AS culture) had'very
low COD and BOD in comparison. The ratio in this last case is of no
particular significance.
We have discussed AS digestion in terms of COD thus far. Because
of the COD:BOD ratios, the performance of AS cultures in BOD terms
can be shown to be markedly better, as follows.
% COD removal = CODp - COD24 * 100 ?
COD0
in a typical case:
757 = 400 - 100 x 100
1 /0 400
With our typical batch feeding regime where four liters of primary
effluent is fed to a starved AS culture with a final volume of ten
liters,
„,_ AS decantate COD x 6 + Primary Effluent COD x 4
COD0 —
in a typical case:
400 = 100 x 6 + 850 x 4
Using the average COD:BOU ratios from Table XIII to convert COD to BOD,
one gets:
BOD0= (100/1.9) 6 + (850/0.72)4 =315 + 4700 = 502
10 10
and
% BOD removal = 502 - 100/1.9 = 90%
502
40
-------
TABLE XIV
COD:BOD Relationship of Various
Waste-Treatment Fractions
WASTE FRACTION
Primary effluent
Activated Sludge
Mixed liquor
Control
COD
910
1010
755
1590
750
750
650
770
800
BOD
1030
1690
1380
1900
520
620
470
350
770
COD:BOD
Ave.
Ave.
1.46
AAB-Fed
Activated Sludge
Decantate
Control
AAB-Fed
Anaerobic Sludge
Digestor Decantate
Control
Iron-rich
725
855
810
90
145
165
135
150
104
144
710
102
95
173
3990
4150
4300
218
165
160
380
520
360
Ave.
55
70
85
5.5
90
60
Ave.
35
800
52
90
52
Ave.
3690
3860
5780
Ave.
210
100
90
1.90
1.65
2.25
1.93
1.65
2.07
1.95
2.45
1.65
1.73
1.92
4.12
0.90
1.96
1.05
3.33
2.44
1.10
1.10
0.75
0.98
1.05
1.65
1.75
Ave.
1.48
41
-------
In the case of AAB-fed AS, with a typical effectiveness of 65% COD
removal and with an AS-decantate COD:BOD ratio of 2.4, the same
calculation shows removal to be also 90%.
Activated Sludge Digestion of Anaerobic Sludge Digestion Decantate
After anaerobic sludge digestors (SD) had operated for about a
month, the whole volumes of the decantates were collected and passed
through activated sludge to determine toxicity (if any) and digestibility.
As is discussed in a later section, and as is shown in Table XIV, the
decantate from the SD fed iron-rich sludge had a low COD (175) whereas
the SD fed control sludge had a decantate with about 4000 mg/1 of COD.
The iron-rich SD decantate was augmented with SNP and fed to a control
AS culture. The decantate from the SD fed control sludge was suitably
diluted and fed to the AAB AS culture. The results are given in Table
XV and Figure 4.
From the results shown, it is obvious that iron-rich anaerobic
decantate is not toxic to ordinary activated sludge, and AAB-fed
AS can handle control anaerobic decantate about as well as it digests
ordinary primary effluent. Other qualities of the anaerobic decantates
are discussed later.
Tertiary Treatment
Activated charcoal adsorption has become a major tertiary treat-
ment method. Zuckerman and Molof (25) recently discussed alkaline
hydrolysis (with lime at pH 11.5) as a preliminary step to adsorption
and they concluded that activated sludge secondary treatment adds
nothing to the over-all effectiveness of treatment when hydrolysis/
adsorption is used for tertiary treatment. In alkaline hydrolysis
treatment it is necessary to readjust pH to the neutral range with
carbon dioxide or strong mineral acid. Although hydrolysis/adsorption
treatment is still in the pilot plant stage of development, it is of
interest to consider how acid brine disposal could affect the method.
Obviously acid brine could be used to adjust pH following alkaline
hydrolysis, and the acid would itself be neutralized in the process.
The hydrous iron oxide precipitate would also no doubt assist in the
flocculation of unhydrolyzed colloidal organic particles and molecules.
On the other hand, the unsettleable hydrous iron oxide might have an
adverse effect on adsorbtion or regeneration of activated charcoal.
Ozone has also been promoted for destruction of refractory organic
matter and for disinfection. Precise pH control and adequate buffering
are very important to the effectiveness of ozonation (26). For the
integrity of activated sludge, it is necessary to neutralize acid brine,
as has been discussed. The neutralization process can be carried to any
pH between 2.0 and 11.5 and the resultant mixture is well buffered.
Copper oxide has been shown to be an important heterogeneous catalyst
in ozonation, but the role of other metal oxide particles (including iron)
is problematical. The impact of lime-neutralized brine on ozonation
is a prime subject for future research.
42
-------
TABLE XV
COD-Depletion Experiment on Anaerobic
Sludge Digester Decantates
AAB-AS
Fed Control Decantate
Time COD %
(hrs) (mg/L) removal
0
2
4
6
24
416
340
265
234
162
43
61
Control-AS
Fed Iron-Rich Decantate
Time COD %
(hrs) (mg/L) removal
0
2
4
6
24
547
416
360
252
108
54
80
43
-------
500
400
300
60
H
§200
u
100
SNF + Fe - rich SD decant
Control SD. decant
12 16
Time, hrs.
Figure A. Activated Sludge Digestion of Anaerobic Decantate
-------
TABLE XVI
Phosphate Removal from Primary Effluent
by Lime and AAB Treatment
Portion Treatment
None
pH = 5.8
Membrane filter
(0.45 aicron)
Add lime to pH 8.5
decant
Add lime to pH 9.7
decant
Add AAB-I 20% vol/vol
Add lime to pH 7.7
Membrane filter
As portion 5, but
decant
As portion 5, but
aerate 15 min. before decant
Total Phosphorus
(mg P/L)
8.50
5.83
3.00
1.45
zero*
trace
trace
* In our method of analysis, zero P means much less
than 0.020 mg P/L; trace means approximately 0,015 mg P/L,
45
-------
Phosphate Removal
Phosphate removal from primary effluent by lime and by AAB plus
lime was studies on a batch of effluent containing 8.5 mg total P/L
and 5.83 mg soluble P/L. Several portions of the effluent were
treated in various ways and the membrane filtrates or decantates
were acid digested and analyzed for P by the standard molybdate-blue
method with the following resul-ts (Table XVI) . From the results it
is apparent that treatment with AAB plus lime gives virtually 100%
removal of phosphate in the filtrate or decantate. In actual plant
practice, where alum, lime, iron salts, or combinations of them are
used for phosphate removal, only slightly more than stoichiometric
quantities are used because of cost, and typically only about 90%
(or less) removal of phosphate is accomplished. For efficient phos-
phorus removal, according to Wuhrmann (20) , the stoichiometric amount
of iron (Fe:P equal to 1.8) has to be supplemented with at least 10 mg
Fe/1 for hydroxide floe formation. Wuhrmann showed the lime sufficient
to raise the pH to 8.8 can be substituted for the excess iron, in which
case sludge quantity and volume will be much reduced and the sludge
produced has better handling properties. If the emphasis is on acid
brine disposal, those points are mott.
The actual amounts of Ca and Fe contained in tests 5, 6, and 7
are shown in Table XVII along with the stoichiometric amounts in
relation to phosphorus based on the empirical formula of the simplest
iron phosphate and calcium phosphate salts. In addition to the iron
and calcium, test protions 5,6, and 7 contained 25 mg Al/1.
TABLE XVII
Fe, Ca, and Al Contents of Test Samples
for Phosphate Removal
Actual Amount Stoichiometric Factor
(mg/L) Amount of
(mg/L) Excess
P 8.5
Fe 200 15.2 13.3x
Ca 325 16.6 19.6x
Al 25 7.4 3.4x
Analyses for nitrogen fractions were not included in the protocol
for phosphate removal test, but some information is available from
the literature. Neil (27) used a combination of alum (94 ppm = 8.5 mg
Al/L) and activated silica (3.4 ppm) for raw sewage treatment and he
reported removal of 98% soluble phosphate; Kjeldahl nitrogen was reduced
46
-------
by 35%. Rohlich (28) observed that 200 ppm of alum (18 rag Al/L)
added to the effluent of the Madison, Wis. Nine Springs Sewage Treatment
Plant reduced phosphate to 0.06 mg P/L (from an initial value in range
5-10 mg P/L), and the treatment effected a 68% removal or organic
nitrogen compounds. Rohlich found no removal or inorganic nitrogen
compounds. There is no reason to expect sorption and/or precipitation
of ammonia or nitrate by lime, alum, or iron salts. Rohlich and
colleagues (29) later reported 60% removal or organic nitrogen (from
secondary effluent) by flocculation with 250 mg alum/L. In the same
tests 25% removal of nitrate and nitrite was found but no removal of
ammonia nitrogen. The possibility of nitrate reduction and/or denitri-
fication by ferrous iron is discussed later. Rand and Nemerow (30)
reported preliminary results of chemical flocculation studies at the
Metropolitan Syracuse Treatment Plant using 5 ppm Fe2(SO,)3 and 30 ppm
lime. For three runs, they observed 48, 64, and 80% phosphate removal.
Kjeldahl nitrogen increased in two of the three runs, which was
attributed to "sampling procedures". On the single run in which
Kjeldahl nitrogen was reduced, 22% removal was found. Thus it is
reasonable to conclude that chemical coagulation of sewage fractions
can remove a significant amount of nitrogen.
Biological Nitrification and Denitrification
It was hoped in the present project to study the effect of AAB
on biological nitrification and denitrification, but we were unable
to produce an adapted sludge culture until the last few days of the
period devoted to experimental work. A description of the manage-
' ment and behavior of the nitrifying sludge culture follows.
—Recipe for feed stock solution (SGP) for nitrifying
Sludge cultures —
Sucrose 15.0 g/L
Gelatin 11.7 g/L (<-o7% N)
KH2P04 0.8 g/L
The mixture has a pH of 6.2
— Ammonium sulfate stock solution (to contain 10 mg
N/M1).
(NH4)2S04 47 g/L
—Soil extract is made from rough turf growing on a fertile
sandy loam low in calcium. Ten grams of air-dried soil is treated
with 100 ml hot water and the mixture is filtered through coarse paper.
47
-------
A typical daily routine would be to feed 15 ml SGP, 10 ml
soil extract (for trace elements) and 2 ml ammonium sulfate (20 mg
N). Sludge solids in the culture were maintained at 1400-1600 mg/l
by wasting or returning settled sludge.
Data on a typical run are given below.
Time Nitrate in decantate
N03 . N (mg/L)
Before feeding 18 pH 5.7
20 min. after feeding 16
40 min, after feeding 4,5
60 min. after feeding 4,0
150 min. after feeding 19
24 hrs. after feeding 20 pH 5,5
On this regime, the nitrate nitrogen content before feeding
will remain in the neighborhood of 20 mg N/L, The diminution of
nitrate after feeding is problematical but we would attribute it
to binding or sorption of nitrate to gelatin and sorption of gelatin
by sludge cells. Metabolic uptake of nitrate is a simpler explana-
tion, but we do not believe that would occur in the presence of 20
mg NH-j.N/L, and under a high rate of aeration. Biological nitrogen
removal has been comprehensively reviewed by Wuhrmann (20) and also
by Balakrishnan and Eckenfelder (31) . According to Wuhrmann the
nitrogen status of sewage fractions is typically as shown in Table
XVIII.
TABLE XVIII
Nitrogen Status of Sewage Fractions (20)
Organic N NH3-N N03/N02-N
Primary effluent 40-45% 50-60% 0-5%
After biological <20% 10-60% 20-80%
Over-a.1.1 proper biological treatment removes about 50% of total
nitrogen from primary effluent. The proportions of ammonia-N and
nitrate/nitrite-N in the effluent from biological treatment varies
over wide limits depending upon operating conditions in the plant.
The peneral scheme for biological nitrogen removal involves digestion
of nitrogenous organic matter by a variety of aerobic heterotrophic
43
-------
microbes, conversion of ammonia to nitrate by Nitrosomonas spp. and
Nitrobacter spp. (nitrification) and conversion of nitrate to
molecular nitrogen (or to N20 according to Wuhrmann) under anaerobic
conditions by a variety of "nitrate-respiring" bacteria- Nitrate can
also be feduced to ammonia by a variety of microbes, and culture
conditions with regard to detention time and loading rate in the
aeration stage must be controlled in order to favor denitrification
over mere nitrate/nitrite reduction (to ammonia) under anaerobiosis.
Nitrification (the oxidation of ammonia to nitrate ) is a
two-step process effected by two raicrobial groups:
Nitrosomonas spp.
1) NH3+ 3/2 02 > N02~+ H20 4- H+
nitrite
Nitrobacter spp.
2) N02~+ 1/202 *• N03~
nitrate
The growth rate of both those groups of organisms is low
in comparison to ordinary heterotrophic bacteria and their growth is
more diminished by low temperatures. Thus in continuous AS culture,
in which both decantate and excess sludge are continuously being
harvested, the nitrifying microbes can easily be washed out of the
culture. In batch culture, the experimentor has more freedom to
alter compositon of the feed, feed dosing, return sludge, etc. but
in the in the long run, the problem is the same, maintaining enough
of the active species to handle the dosing of ammonia. Over-all,
nitrification is dependent upon conditions of operation which
minimize excess sludge production. The critical oxygen tension for
nitrification is lower than formerly thought. A concentration as
low as 1.0-1.5 mg 02/L is adequate.
As we mentioned, denitrification requires anaerobic conditions;
this condition can be achieved simply by holding sludge without air
for a time. After a period of aeration and limited feeding sufficient
to induce nitrification, exogenous substrate has been used up but
sufficient endogenous substrate remains to maintain the nitrate
respiration of the sludge until all the nitrate is used up. It is
not necessary to add raw sewage or other exogenous sources of hydrogen
donor substances if endogenous reserves are present in amounts
equal to the nitrate/nitrite. In fact one gets the impression that
exogenous substrate may promote nitrate reduction as opposed to
denitrification,
49
-------
Biological nitrogen removal can thus be seen to depend upon
a number of parameters, some hydraulic and some biological or chemical.
In the disposal of acid brines through municipal waste treatment plants,
nitrogen removal could be profoundly affected; for example, high
ferric iron concentrations may favor nitrifying species over ordinary
heterotrophs, ferrous iron can reduce the time needed for nitrifying
sludge to pass into anaerobiosis in the denitrification vessel. These
questions are amenable to experimental inquiry and they should be
given prompt attemtion in the future.
Anaerobic Sludge Digestion
The digesters were operated for two months, during which time
they received about thirty-five 100-ml portions of settled sludge.
Volume was maintained constant by dedecanting off liquor after settling
the sludge in the digesters. Table XIX gives the time course of
build-up of solids in the digesters.
Over the two-month period, the mixed liquor suspended solids
built up from 8.2 to 9.3 gm/L in the control digestor. In the digester
fed iron-rich sludge, the solids built up from 14.8 to 20,1 gm/L.
Both digesters were originally charged a month previous to the period
reported with 400 ml of heavy digestor sludge plus 2500 ml of water,
the solids level in both digesters being initially about 7,0 gm/L.
During the period reported in Table XIX, the control digestor received
about 36 grams dry weight of sludge, and the iron-rich digestor
received about 40 grams. Table XX summarized the sludge digestion
data.
The specific digestion rate is of special interest; it should
be noted that it is independent of volume and sludge concentration,
The specific digestion rate designates the fraction of resident
sludge that can be digested in a day. In other words, about 2% a
day in the control and slightly less than 1% in the iron-rich digestor.
Anaerobic Sludge Digestor Decantate
We have already discussed the COD and BOD quality of the
decantates. The phosphate and iron contents of the decantates are
given in Table XXI.
The difference in phosphorus content between the control and
the iron-rich digestor decantates is quite striking, while the
relatively small difference in iron content is surprising. We
hasten to point out that even though analyses were run on decantates
and filtrates, it would be incorrect to conclude that the contents
shown represent iron and phosphate in solution. Decantates (actually
coarse-paper filtrates in some cases) are quite turbid and the analytical
methods used release iron and phosphate into solution from particles
in the sample. The control liquor filters extremely poorly; in fact,
50
-------
TABLE XIX
Anaerobic Sludge Digestion in Digesters
Fed 100 ml Sludge Approximately Daily
Elapsed Time Mixed Liquor Suspended Solids
(days) (gm/L)
0
3
7
10
14
16
21
24
27
29
33
36
41
46
49
53
56
59
Control
7.8
7.6
7.4
10.0
8.2
8.4
9,6
8.4
9.1
9.0
8.0
8.6
8.2
9.0
8.2
9.8
9.0
10.0
9.0
—
9.5
9.0
Iron-rich
14.8
13.8
14.2
15.0
14
10.8
15.0
14.0
13
15.6
15.2
15.8
15
15.0
17,2
17.8
16
16.2
18.6
17
20.6
19.6
,5
.3
.5
.7
.4
9.3 20.1
The table shows day-by-day values and averages over 10-day periods.
51
-------
TABLE XX
Summary of Anaerobic Sludge Digestion Data
Control Iron-Rich
I Range of dry suspended solids 0.5-2,1 0.4-2,2
fed (gm/100 ml)
II Average of dry suspended solids 1.04 1.15
fed (gm/100 ml)
III Total for 35 100 ml portions 36.4 40,2
(gm)
IV Average daily feed (111/60) 0.60 0.67
(gm/day)
V Initial total resident dry sludge 23.8 42.0
VI Final total resident dry sludge 27.0 58,3
(gm)
VII Sludge increase in 60 days (gn) 3,2 16.3
(VI-V)
VIII Average daily increase in dry 0,05 0.27
sludge (VII/60) (gm/day)
IX Average daily sludge digestion 0,55 0,40
(IV-VIII) (gm/day)
X Specific digestion rate (gm/day. 0,022 0.008
gm (IX/ (VI + V) /2)
it doesn't centrifuge very well:. The iron-rich liquor both filters
and centrifuges remarkably well* " .iwer phosphorus content of
iron-rich decantate can thus be uted partly to higher iron and
partly to improved filterability. 'Another extremely striking difference
between the two digesters is odor. The control digester odor is about
as foul as can be imagined, typical of an anaerobic sludge digester.
The iron-rich digester has a mild odor reminiscent of crude petroleum.
Digested Sludge Disposal
Time did not permit any experiments on sludge drying or
incineration, but samples of both control and iron-rich sludge have
been frozen and saved for future work. Certain things can be said,
52
-------
however, even without experimental evidence. First, the iron-rich
digested sludge is no doubt largely a mixture of ferrous oxides and
ferrous sulfide probably amorphous in form rather than crystalline,
TABLE XXI
Characteristics of.Anaerobic Sludge
Digestor Decantate
Control
Range Average
Phosphorus (mg/L) 8-92 57.6
Iron (mg/L) 22-69 47.8
COD (mg/L) - 4150
Iron-Rich
Range Average
Phosphorus (mg/L) 2^15 6.0
Iron (mg/L) 31-222 118.0
COD (mg/L) - 181
In the course of anaerobic digestion sulfates become reduced
to sulfides and the sulfhydryl compounds of digested microbial
tissue also suffer the same fate.
The land disposal of sludges rich in iron sulfide could pose
a problem unless full assurance against oxidation could be provided.
Oxidation and leaching of such a sludge dump would produce a leach
water similar to acid mine drainage. Incineration of the sludge
would produce a quantity of sulfur dioxide, and stack scrubbings
would also resemble acid mine drainage. The iron-rich sludge is
much higher in iron relative to sulfur than are pyrites, the
minerals which produce acid mine water in weathering. In the
disposal of acid brine in municipal waste treatment plants, the iron
will pass very largely into sludge while the sulfate will move
through the plant in decant fractions, The ratio of iron to sulfur
in digested sludge should be determined in future work. In any case,
digested iron-rich sludge probably has about the same composition
as limonite iron ores and as such should be of some, albeit limited,
value,
53
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Sludge Conditioning
All methods of sludge disposal are benefited by conditioning
or dewatering. Chemical conditioning of sludge by sulfuric acid,
aluminum sulfate, ferric sulfate, ferric chloride, and lime was
reviewed by Balakrishnan et al (32). Relative dewatering rates
more than 100 times the untreated rate have been achieved through
the use of the chemicals listed above. We have alluded to the improved
filterability by primary sludge, activated sludge, and digested
mixed sludge from sewage treatment using lime-neutralized acid brine.
Masselli et al (33) have reported that 2000-5000 gal. primary sludge
is produced per million gallons of raw sewage and 5000-15,000 gal.
activated sludge are produced from the same quantity of primary
effluent. Chemical conditioning of sludge uses chemicals in the
range of a few hundred parts per million, and thus only a small part
of the quantity of acid brine available for use at a municipal waste-
water treatment plant would be used in sludge conditioning, but there
is no doubt about the potential value of acid brine in sludge
conditioning. If acid brine were introduced at the primary stage of
sewage treatment, there might not be any need for the sludge condition-
ing step, as we have mentioned,
Oxidation of Ferrous Iron in AAB by Oxygen
In order to determine the additional aeration capacity needed
for the oxidation of ferrous iron in acid brines added to sewage
and to determine the rate of the reaction under typical conditions
two simple experiments were carried out. In the first, samples of
AAB neutralized with lime to pH 5.25 were added to tap water samples
saturated with dissovled oxygen. Dissolved oxygen was measured with
a YSI model 54 oxygen meter at the beginning of the tests, after 10
minutes, and after 15 minutes. The test jars were left open to the
air but were not shaken or stirred. The results are given in Table
XXII.
TABLE XXII
Reduction of Dissolved Oxygen by Ferrous Iron
Test # Ratio of Fe-H- to 02 % of 02 used up in
present (equivalents) 10 Min. 15 Min.
1 0.9 22 22
2 1.9 — 43
3 3.0 53 57
4 4.2 — 73
5 5.6 70 83
54
-------
The data in Table XXII can be interpolated to conclude that even
under quiescent conditions the dissolved oxygen will be 50% reacted
in 5-10 minutes at high oxygen demand levels due to ferrous iron.
The actual reaction rate depends upon the concentrations of the
reactants and the amount of mixing. At a feed rate of 20% AAB
(1000 rag Fe-H-/liter), the zero-time ferrous iron concentration is
200 Fe-H- per liter; for oxidation to be complete in a ten-hour
period, 20 mg Fe++/L/hr must be oxidized requiring theoretically
20/7 = 3.0 mg 02/L/hr. For a four-hour detention time, the oxygen
rate requirement would be 7.0 mg 02/L/hr. With typical activated
sludge plants the oxygen supply rate from aeration is more than 100
mg 02/L/hr.
In a second type of experiment, a two-liter sample of AAB-I
was neutralized with lime and added to an eight-liter batch of
freshly fed activated sludge. The mixture had a pH of 6.8, The
mixture was then aerated in our standard AS set-up. Small samples
of the culture were taken at intervals, filtered through a membrane,
and the filtrate was analyzed for iron. At pH 6,8, it can be
assumed that only ferrous iron would remain in solution. The change
"in soluble iron content with time is shown in Table XXIII and
Figure 5.
TABLE XXIII
Change in Time of Soluble Iron in AS containing
20% Neutralized AAB
Time (after neutralizing) Fe in Filtrate
(hours) (mg/L)
0 50
2 6,3
4 3.8
6 1.0
It should be noted that the ferrous iron content of the mixture was
initially 200 mg/L; thus, three-fourths of the ferrous iron was
oxidized and precipitated or was scavenged by the precipitated
ferric hydroxide floe in the few minutes required for neutralizing
the MB, sampling the AS mixture, and filtering the sample. From
the color change in the AS mixture from greenish brown to orange
brown almost all of the ferrous iron added was oxidized (and
precipitated) in the first few minutes of aeration.
The aeration requirements in engineering terms are given in Table
XXIV and Figure 6.
55
-------
50
Ui £
O 4J
•H
C
40 .
- 30 .
20
10
0
0
1. Initial Concentration 400 mg/1 (200 rag/1
and 200 mg/1 Fe+3)
2. 3/4 of Ferrous iron was precipitated in a
few minutes
1234 56
Time, hrs.
Figure 5. Disappearance of Soluble Iron in AS Containing AAB Neutralized with Lime
-------
TABLE XXIV
Oxygen and Work Requirements for Oxidation
of Ferrous Iron in Acid Brine
Fe-H- Increased D.O. Increased Work
(mg/L) Requirement (mg/L) Requirement H,P. Hr,
5 0.7 0.125
10 1.4 0.25
25 3.5 0.625
50 7.0 1.25
100 14.0 2.5
200 28.0 5.0
500 70.0 12.5
Assumptions:
1. Stoichiometric relation is one equivalent oxygen per
equivalent iron. In English units: one pound of oxygen oxidizes
seven pounds of ferrous iron.
2. Aeration efficiency is 2//02/H,P.Hr.
3. 1.0 mgd capacity.
Actual aeration requirements would be slightly, but only slightly,
higher because of the speed of the oxidation reaction in relation
to detention times in actual practice.
Microbiological Status of AAB AS Cultures
The term iron bacteria has been used variously to describe
microbes associated with iron oxide crusts or deposits in water,
soils, and in man-made systems. In the general sense in which the
term, iron bacteria, is used in Standard Methods, both autotrophic
and heterotrophic forms are included. In the former instance
carbon dioxide is fixed with energy derived from oxidation of ferrous
iron, and in the latter instance, some form of organic carbon is
assimilated. Heterotrophic bacteria can also bring about the precipi-
tation of hydrous iron oxide by digesting the organic ligand from
soluble ferric chelates, or the surfact of bacterial colonies can
simply serve as a nucleus for the precipitation of ferric hydroxide
57
-------
oo
Assumptions:
Stoichiometric relation is one equivalent
oxygen per equivalent iron;
ie 0,14 mg/1 02 is needed to convert
1 mg/I Fe+2 to Fe+3
Aeration Transfer = 2 Ib.
Hr.
0
400
450
200 250 300 350
Fe+2 Concentration, mg/1
Figure 6. Theoretical Oxygen Requirements For Oxidation of Ferrous Iron
17.5
. 15.0
12.5
..10.0
7.5
5.0
.. 2.5
500
3
O
i-l
n>
ca
05
n>
D-
ro
c
H-
H
rt
01
-------
formed by chemical oxidation of ferrous iron in solution. Metallic
iron and pyrites (FeS2) are attacked by natural waters, especially if
the water contain dissolved oxygen. Iron bacteria are often if not
always associated with the dissolution or corrosion process in nature.
The special role of iron (and sulfur) bacteria in the formation of
acid mine waters from pyrite was very recently elucidated by Baker
and Wilshire (34). Those authors pointed out that Ferrobacillus
ferrooxidans oxidizes only ferrous iron, Ferrobacillus sulfooxidans
oxidizes S= and ferrous iron, and Thiobacillus thiooxidans oxidizes
only S= (and thiosulfate). All three of these autotrophic bacteria
are capable of growing in media with a pH as low as 2.0, but that
capability has not been shown for the some dozen of other species
described as "iron bacteria",
Pringsheim (35,36,37) demonstrated that the "sewage fungus"
Spaerotilus natans can grow as autotrophically as an iron bacterium,
and Skerman et_al_ (38) demonstrated that species' ability to grow
as a "sulfur bacterium"} thus it seems to have the same range of
chemoautotrophic capability as Ferrobacillus sulfooxidans. The
taxonomic status of many microbial entities, described as iron
bacteria, is problematical as witnessed by Lundgren's (39,40,41)
reference to Thiobacillus ferrooxidans as a participant in acid mine
water genesis.
Baker and Wilshire in their studies on pyrite dissolution and
oxidation found a variety of ordinary aerobic heterotrophs in
their reactors, and they concluded that these microbes, (Penicillium,
Pseudomonas, Aerobacter) growing on organic moieties contained in the
coal particles associated with the pyrites, play a significant role
in the genesis'of acid mine water by supplying carbon dioxide to the
autotrophic iron bacteria; at low pH carbon dioxide is far less
soluble than it is near neutrality.
In summary, chemoautotrophy in iron bacteria, postulated by
Winogradsky in 1922, was unproven in 1949; a large number of
filamentous iron bacteria are referrable to Sphaerotilus natans,
an obligate aerobic variable species which grows perfectly well
heterotrophically but does not tolerate low pH. At pH close to
neutral, ferrous iron is oxidized so fast chemically by dissolved
oxygen that it is doubtful that microbes can avail themselves of the
energy. Pringsheim did not rule out the possibility, however, and
he pointed out especially that Sphaerotilus in nature grows where
ferrous iron, often complexed with humates, is to be found in oxygenated
water. At pH 2-3 ferrous iron is oxidized very slowly by dissolved oxygen
except in the presence of a bacterial catalyst. Chemoautotrophy in
Ferrobacillus sp., which grows at pH 2-3, was convincingly demonstrated
by Lundgren and colleagues in the fifties, and the point was further
amplified by them in several papers in the sixties (14,15,42,43).
59
-------
Our own work on the role of iron bacteria in activated sludge
fed AAB consisted of attempts to isolate Ferrobacillus from the AS
cultures using Silverman & Lundgren's medium (15) and following
procedures outlined by them. Both control AS cultures and AAB-AS
cultures yielded a mixture of heterotrophs, including pleomorphic
molds, and iron bacteria in Silverman & Lundgren's medium. The
granular and crusty deposits formed in the iron bacteria cultures
was in every way reminiscent of the material called "yellow-boy"
in acid mine waters. Our experience with small iron-bacteria cultures
using an acid medium containing ferrous iron correspond very well
with those of Baker & Wilshire (34), who used a continuously flowing
pilot plant with pyrite as the reduced iron source. In their case,
the growth of heterotrophic microbes was supported by coal substances
and in our case we are forced to conclude that the substrates were
substances leached from new polyethlene stock solution bottles—the
medium we used contained no organic substrates in its recipe.
As far as the identification or enumeration of Ferrobacillus
S£p_ is concerned, we have not succeeded in obtaining growth on
the solid medium recommended for that purpose by Lundgren & Schnaitman
(42) from AS cultures or from the inorganic liquid enrichment cultures
derived from them. It is quite clear, however, that inoculated
liquid cultures form yellow-boy in a few days and uninoculated flasks
do not. In any case, this discussion of the physiology and ecology
of iron bacteria is academic in the present context; activated sludge
at neutral pH does not provide the environment for the growth of
chemoautotrophic iron bacteria, even though they obviously survive
in the environment. Since that group of microorganisms grows
autotrophically by the oxidation of ferrous iron, sulfur or siilfide,
there is no reason to believe that they should be important in the
economy of AS.
Wuhram (20) pointed out another, more important, microbiological
consequence of high iron in activated sludge. He observed that iron
in excess of 10 mg/L caused the complete disappearance of the
protozoan fauna within two days. In his work, technical grade iron
salts were used; thus, toxicity from contaminating heavy metals
cannot be ruled out as the cause of the disappearance. Our control AS
cultures characteristically contained three types of protozoa, small
fast-moving flagellates, small and medium-sized notile ciliates, and
large flash-shaped stalked ciliates (vorticellids). AAB-fed AS
cultures contained no protozoa except for a problematical large
subspherical "amoeba" x^ith a very dense granular protoplasm, blunt
pseudopodia, and a conspicuous rind or plasmalemma. These "amoeba"
change shape very slowly (if at all); these objects may be immature
mold sporangia, but no mycelial strands were evident in the culture.
Protozoa in AS play a manifold role. Through predation they
remove bacterial strains that might otherwise build up to the
detriment of the system. In addition they are responsible for the
digestion o£ a part of the organic load and the assist in flocculation,
60
-------
The latter function is probably exercised by small, dense excretory
particles from the protozoa acting as "cement bridges" between
larger, less-dense particles, mostly clumps of bacteria. The loss
of protozoa in AAB-fed AS need not be particularly detrimental,
however, because of the physico-chemical flocculative action of the
iron hydroxide particles, but the absence of protozoa could have a
long-term detrimental effect because of the lack of cropping of
undesirable bacterial strains.
Concentration of AMD by Reverse Osmosis
Reverse osmosis has been chosen as the process of concentrating
acid mine drainage effluent to acid mine drainage effluent to acid
brine because of the following reasons:
1, Process is feasible at reasonably economic levels,
2. The potential for considerable reduction in the cost of
of concentration is high due to the anticipated improve-
ment in the membrane science and technology.
The flow chart for the concentration process is shown in Figures
7 and 8. The total dissolved solids (TDS) level of the acid mine
drainage effluent is assumed at 50 mg/1. The TDS level of concentrated
acid brine at the end of two stages is 675 mg/1 and concentration
factor is about 13. With the addition of third stage reverse osmosis
unit, the TDS of concentrated acid mine will be about 2400 mg/1
resulting in a concentration factor of close to 50,
The unit cost figures were taken from the information published
by Channabasappa and Harris in Industrial Water Engineering (44),
These cost figures were adjusted for second and third stage reverse
osmosis units to allow for the type of membranes to be used, pH
adjustment etc. The costs of concentrating acid mine drainage
effluent to 675 mg/1 TDS and 2400 mg/1 TDS are calculated for plant
sizes 1 and 10 MGD as follows:
1 .MGL Feed
Cost for the first stage is assumed at 55C/1000 gallons product
water. Allowing 10% and 50% increase for the second and third stage
respectively over the first state the unit costs for the second stage
is 61^/1000 gallons product water and 83C/1000 gallons product water
for third stage.
(1) To achieve a concentration of 675 mg/1, two stages of
R.O. units have to be employed. The cost for the two
stages:
61
-------
1 MGD @ 50 ing/I
500 psi
0.25 MGD @ 188 mg/1
to
0.75 MGD
@ 7 mg/1
First Stase
1
0.0625 MGD @ 675 mg/1
0.1875 MGD
@ 25 ng/1
Second Stage
0.0156 MGD <§
,430 tng/1
0.0469 MGD
@ 90 mg/1
Third Stage
rlgure 7. Flow Diagram for Concentrating of AMD - one MGD Plant
62
-------
10 MGD @ 50 rag/I IDS
500 psi
2.5 MGD @ 188 mg/1
7.5 MGD
@ 7 mg/1
First Stage
1
500 psi
0.625 MGD (3 675 mg/1
0.156 MGD
,Ws
1.875 MGD
-------
= $0.55 (750.000) + $0.61 (187,500)
1000 gal 1000 gal
= 413 + 114
= $527 / 62,500 gallon of concentrated AMD.
Cost of producing 1000 gallons
of concentrated AMD @ 675 mg/1 = $527 (1000)
62,500
= $8.45
(2) For the concentration of acid brine to 2400 rag/1, three
stages have to be employed. The total cost for all the
stages:
- $0.55 (750,000) + $0.61 (187,500) +
1000 gal 1000 gal
$0.83 (46,900)
1000 gal
= 413 + 114 + 39
= $566/15,600 gallon of concentrated AMD
Cost of producing 1000 gallons of
concentrated AMD @ 2400 mg/1 = $566 (1000 )
15,600
= $36.4
Cost of Transporting by Tank Truck
Source: 1. McCormack T.I, Trucking Co., Inc.
Bulk Liquid Transporters
U.S. Highway No. 9
Woodbridge, New Jersey
201-382-4800
2. MatLac Pennsylvania - 215-CL9-9800
Trucking Charges up to 100 miles
23.5 cents/100 Ib on a minimun of 40,000 Ib.
27 cents/100 Ib for 40,000 Ib.
Cost of transporting by tank trucks
1000 gallons = 23.5_ 1 (8340)
100 ' 100
= $19.60
64
-------
Transporting by Rail
Source: Reading Pennsylvania Railways
212-349-2175
For distances less than 10 miles,
$0.58 (8340 Ib) = $48.5/1000 gallons
100 Ib
For distances more than 10 miles
$0.49 (8340 Ib) = $41,00/100 gallons
100 Ib
For bulk transportation and an extended period,
special reduced rates could be obtained by writing
to the railways.
Cost of Transporting Brine via Pipeline
The volume of concentrated brine considered for transporting
is in the range of 10,000 gpd to 250,000 gpd. The distance to be
transported is in the range of 10 to 50 miles. The cost figures for
the four cases are determined as below:
Case 1
10,000 gpd - 50 miles ,
From a preliminary engineering analysis, a 4"(7) reinforced
plastic pipe and 8 hours pumping are selected,
A. 50 miles, 4"G> reinforced plastic pipe
(? $1.40/ft, = (5280) (50) (1.40) = $268,000
B. Cost of pumphouse 5,000
C, Cost of pumps and motors 2,000
D. Excavation and filling 3' x 3'
section assumed. Volume =
(5280) (50) (3 x 3) /27 = 88,000 cu, yd.
@ 55 cu. yd/hr., hrs, of backhoe operation
= 38?000 = 1600 hrs. or 200 days
55
Backhoe and operator 0 $320/day 200 x 320 = 64,000
2 laborers @ $50/day x 200 days = 20.000
84,000
34,000
Total Capital Cost 359.000
Contingencies (? 15% 54.000
413,000
Engineering 9 10% 41^300
il4,300 or 455,000
65
-------
Amortizing the capital cost for 20 yr life and 5%
interest, annuity cost = 455,000 x (8,024) (10~2)
= 36,500
Daily cost = 36f500 = $100.00
365
Power: 5HP @ 1 p 24 hr/day = 0.50
HP hr,
Maintenance @ 2 hrs. operators )
time/day at $100/day) = 25-°°
Total operating Cost 125.50
~-.
-------
Case 3
250,000 gpd - 50 miles ,
Pipe size selected = 8"Q reinforced plastic pipe
Pumping hours = 24 hrs.
HP =25
A, 50 miles, 8"© reinforced plastic pipe
-------
Subtotal 309,800
Contingencies @15% 46,200
356,000
Engineering @ 10% 35,600
Total Capital $ 391,600
Cost
Amortizing for 20 yrs 5% interest
annuity cost = 391 600 (8.024 x 10~2) = 86.00
365
Power 2.00
Maintenance as is Case 3 100,00
Total operating Cost $188.00
Cost of pumping 1000 gallons ) _
of acid brine - 10 miles ) = i§§ (100°)
250,000
= $0.75
The costs of pipelining are shown in Table XXV. In that table
the numerator figures are the costs per thousand gallons for the
distances shown, and the denominator figures are the cost per thousand
gallons per mile for the plant sizes shown,
TABLE XXV
Pipeline Transport Cost Estimates
Plant Size Distance
10 miles 50 miles
10,000 gpd $4.71 12.55
T574T "TT.TT
250,000 gpd 0.75 2.15
07043
68
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SECTION VIII
SUMMARY
This report examines the feasibility of concentrating acid mine
waters by certain processes and transporting the concentrated brine
to municipal wastewater treatment plants for disposal. Relatively
pure water is a by-product of the acid mine water treatment, and the
iron-rich acid brine is of some value in sewage treatment. The
artificial brine used in the laboratory research was devised to
simulate a fifty-fold concentration of acid mine water with respect
to iron and acidity, (2000-10,000 mg Fe/L and acidity 7,800-23,000
ing/L as
In experiments on primary settling acid brine was added at levels
up to 50% vol/vol with no significant effect on the process. With
acid brine neutralized with lime, a marked improvement in primary
settling was found. In addition, virtually 100% removal of phosphate
is achieved,
Activated sludge digestion was found to be completely inhibited
by acid brine at the levels included in this study, but lime-
neutralized brine was not particularly inhibitory. Over a two-month
period the control activated sludge culture achieved as high as 89%
removal of COD and averaged 71% removal. The activated sludge culture
fed 20% vol/vol lime-neutralized acid brine achieved as high as 90%
COD removal and average 64%. The corresponding averages for BOD
removal were 90% in both the control and the acid brine activated sludge
units.
Anaerobic sludge digestion was studied using the daily harvests
from control and acid brine-fed activated sludge units over a two-
month period. The specific digestion rate in the control digester was
2% per day and 1% per day in the digestor fed iron-rich sludge. Both
digesters produced a decantate that was readily digested by activated
sludge. In contrast to the control anaerobic digestor, the iron-rich
digestor produced a decantate that was low in phosphate, low in COD,
and virtually odor-free.
Oxygen and work requirements for oxidation of ferrous iron in
acid brine were determined to be 0.14 mg oxygen per milligram ferrous
iron and 0.025 horse power hour per milligram ferrous iron in a one-
mgd plant. Thus in a one-mgd plant handling acid brine and having 500
mg/L ferrous iron in the plant flow, the work requirement^ for pumping
air to oxidize the ferrous iron is 12.5 horse power hour.
The microbiological status of the acid brine activated sludge
unit was monitored along with the control, Activated sludge at
neutral pH does not provide a suitable environment for the growth of
chemoautotrophic iron bacteria, but they can be isolated from such
69
-------
cultures. As had been previously observed in Switzerland where iron
salts are used for nutrient removal in sewage treatment, protozoa
were absent from our acid brine-fed activated sludge unit. Their
absence, however, did not cause any impairment of flocculation in the
sludge.
Reverse osmosis offers an attractive method for concentrating
acid mine water. The process produces a concentrated brine and a
relatively pure water as a by-product, A three-stage system could
produce a brine of 24000 mg TDS/L from acid mine water containing 50
mg TDS/L - a concentration of about fifty-fold. The cost of brine
production in a 1-mgd plant is estimated to be $36,40/1000 gal, brine
or 76C/1000 gal. acid mine water treated. The process will produce
47,000 gal. pure water for every 1000 gal. of brine produced.
Pipeline transport of the brine from the mine site to municipal
sewage treatment plants is estimated to be cheaper than truck or
rail, but the pipeline method requires a capital outlay, right-of-
way purchase, and construction, Tanktruck transportation costs would
be on the order of $20,00 per 1000 gal. for up to 100 miles. Standard
rail rates would be roughly twice that amount, but bulk rates and an
extended contract would permit a reduction of the rate. The pipeline
transport costs, not including costs for right-of-way but including
amortization of other capital outlay, are estimated to be in the
range of $4.30 per thousand gallons for a one-hundred mile pipeline
of 250,000 gpd capacity.
70
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SECTION IX
ACKNOWLEDGEMENTS
This report was written by R.J. Benoit, Ph.D and S, Balakrishnan,
Ph.D. It describes work under Contract No. 14-12-847 between the
Federal Water Pollution Control Administration and Environmental
Research & Applications, Inc. of Wilton, Connecticut. Dr. J.M. Shackelford
was project officer for FWPCA, The author gratefully acknowledge the
assistance of Dr. M. H, Naimie and Mr. A. G. Attwater in the laboratory.
71
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SECTION X
REFERENCES
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3. Gerard, L., and Kaplan, R, A,, "Design and Economics of Acid
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73
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14. Lundgren, D. G., and Schnaitraan, C.A., "The Iron Oxidizing
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15. Silverman, M.P., and Lundgren, D. G., "Studies on the Chemoautotrophic
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27. Neil, J.H., "Problems and Control of Unnatural Fertilization of
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28. Rohlich, G. A,, "Chemical Methods for the Removal of Nitrogen and
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A. Taft Sanitary Eng. Ctr., Cincinnati, Ohio, 1961, p. 130-135
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from Secondary Effluent by Alum Flocculation and Lime Precipitation,"
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I. "Nitrification in the Activated Sludge Process"
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the Effects of Industrial Wastes on Sewage Treatment." (TR-15),
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34. Baker, R.A., and Wilshire, A.G. , "Microbiological Factor in Acid
Mine Drainage Formation: A Pilot Plant Study" Env. Sci. & Tech.
4: (5) 401-407, (1970).
35. Pringsheim, E.G., "Iron Bacteria," Biol. Rev. (Cambridge) 24:
200-245 (1949).
36. ibid. "The Filamentous Bacteria Sphaerotilus . Leptothrix.
and Their Relationship to Iron and Manganese," Phil. Trans.
Roy. Soc. London. Ser. B. 233;453-465. (1944) .
37. ibid. "Iron Organisms," End£amn 11"208-214. (1949).
38. Skerman, V.B.D., Dement jeva, G., and Carey, B.J., "Intracellular
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(1957).
75
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39. Agate, A.D., Lundgren, D. G., and Vishniac, W., "Control of
Thiobacillus ferrooxidans from Acid Mine Water," Abstract of
paper presented at the 3rd Int'l. Conf. on Global Impacts of Appl.
Microbiol., Bombay., (Dec. 1969).
40. Silverman, M.P., and Ehrlich, H.L., "Microbial Formation and
Degradation of Minerals," Adv. in Appl. Microb. 6:153-206, (1964).
41. "Sulfide to Sulfate Reaction Mechanism," USDI FWPCA Report 14010
FPS 02/70 USGPO, Research Fdn, Wash. D.C., Ohio State,, (1970).
42. Maclay, W.G., and Lundgren, D.G., "Carbon Dioxide Fixation in the
Chemoautotroph, Ferrobacillus ferrooxidans," Biochem. and Biophys.
Res. Comm. 17;603-607, (1964).
43. Dugen, P.R., and Lundgren, D.G., "Energy Supply for the Chemoauto-
trophic, Ferrobacillus ferrooxidans." J.Bact. 89:825-834, (1965).
44. Channabasappa, K.C., and Harris, F. L., "Economics of Large-Scale
Reverse Osmosis Plants," Ind. Water Engrng, 20-24 (Oct., 1970).
76
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Stibjeft Field & Group
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INPUT TRANSACTION FORM
Organization
Environmental Research & Applications, Inc.
24 Danbury Road
Wilton. Connecticut 06897
Title
Concentrated Mine Drainage Disposal Into Sewage Treatment Systems
10
Authors)
Benoit, R.J.
Balakrishnan , S .
Attwafrer, A.J.
l,t Project Designation
" Program # 14010 FBZ,
21 Note
Contract # 14-12-897
22
Citation
23
Descriptors (Starred First)
* Effects of Soluble Iron Waste, ^Municipal Waste Treatment Facilities, Primary
Settling, Activated Sludge Process, Sludge Digestion, *Sludge Conditioning,
Biological Nitrification, *Biological Denitrification, * Phosphate Removal,
Concentration of Acid Mine Drainage,
25
Identifiers*(Starred First)
*Effects of Soluble Iron, Municipal Waste Treatment
27
Abstcac.
Laboratory scale studies were carried out to evaluate the disposal methods for
concentrated acid mine drainage in the municipal waste treatment. The studies
indicate 'that the concentrated acid mine drainage produced, are not of sufficient
value to pay for treatment by any known method, but they can be disposed of through
municipal wastewater treatment plants, where they will be of some value especially
in sludge conditioning and nutrient removal.
Abstractor
Balakrishnan
Inxt
tution
Environmental
Research
&
Applications,
Inc.
WR:I02 (REV. JUt-Y 19S9)
VRSIC
SEND TO: WATER RESOURCES SCIENTIFIC INFORMATION CENTER
U.S. DEPARTMENT OF THE INTERIOR
WASHINGTON. D. C. 20240
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