Water Quality
T1 Introduction
1. Surface Water Quality Management
In an idealized setting, a community development will permit use of a watershed for
drinking water supply, for recreation and for the discharge of treated wastes.
T2 Watershed and beneficial uses [ask class to identify locations for each]
T3 Watershed and beneficial uses – best case solution
T4 Everyone lives downstream
T5 Everyone lives downstream – Sonora River
2. River Water Quality
T6 Discharges to rivers
T7 Engineering design to protect downstream stakeholders
Here we will consider the management of river oxygen resources as impacted by the
discharge of oxygen demanding wastes, i.e. BOD. The task is best summarized by
considering the battle between oxygen consumption (BOD exertion) and oxygen
replenishment (reaeration).
River DO-BOD Overview
First consider the fate of BOD (which behaves as a first-order decay) in a river after it
exits the mixing basin.
T8 First order BOD decay
Next consider the compensating effect of reaeration, the rate of which increases as the
oxygen concentration decreases. The net effect is a profile with distance downstream in
the river termed the dissolved oxygen sag curve.
T9 The DO Mass Balance and the DO Sag Curve (the battle for oxygen)
dO2
 ka  D  k L  L
dt
Deoxygenation: the BOD Portion of the DO Mass Balance
Aerobic organisms derive the energy required for maintenance of metabolic function,
growth, and reproduction through the process of respiration.
Chemoheterotrophs are organisms that utilize organic matter (i.e., C(H2O)) as a carbon
and energy source and, under aerobic conditions, consume oxygen in obtaining that
energy,
C  H2O  O2  CO2  H2O  
Chemoautotrophs are organisms that utilize CO2 as a carbon source and inorganic matter
as an energy source, and usually consume oxygen. For example, in nitrification, the
microbial conversion of ammonia to nitrate, ammonia provides the energy and
bicarbonate ion contributes carbon,
NH +4  2 HCO3  2O2  NO3  2CO2  3H 2 O  
Because oxygen is consumed in both of these microbially-mediated reactions,
biochemical oxygen demand (BOD) can be defined as the amount of oxygen utilized in
performing the oxidation. And it follows from this that BOD is a measure of the
“strength” of a water or wastewater: the greater the concentration of ammonia-nitrogen or
degradable organic carbon, the higher the BOD. Note that chemical strength (mg
C(H2O)/L or mg NH3-N/L) is expressed here in terms of its impact on the environment
(oxygen consumed, mg BOD/L), i.e. representation by effect.
The simple carbohydrates produced through photosynthesis are used by plants and
animals to synthesize more complex carbon-based chemicals such as sugars and fats.
These compounds are utilized by organisms as an energy source, exerting a
carbonaceous oxygen demand (CBOD). In addition, plants utilize ammonia to produce
proteins, i.e. complex, carbon-based chemicals with amino groups (-NH2) as part of their
structure. Proteins are ultimately broken down (proteolysis) to peptides and then amino
acids. The process of deamination then further breaks down the amino acids, yielding a
carbon skeleton (CBOD) and an amino group. Conversion of the amino group to
ammonia (ammonification) completes the degradation process. The ammonia is then
available to exert a nitrogenous oxygen demand (NBOD) when utilized by
microorganisms.
T10 Sources of BOD
The oxygen requirement for complete oxidation of ammonia-nitrogen to nitrate-nitrogen
or a carbon-based compound to carbon dioxide and water is termed the theoretical
oxygen demand (ThOD). For ammonia, the observed oxygen demand is equal to the
ThOD because all of the ammonia is oxidized. Carbonaceous compounds, on the other
hand, are not all easily or completely oxidized by microorganisms (biodegradation) and
the rate of that oxidation may vary widely among different sources of organic matter. For
example, the carbon-containing compounds present in a Styrofoam cup are not as
biodegradable as those found in tree leaves, and both are less biodegradable than the
carbon compound which constitutes sugar. Thus, while a mass of carbon present as sugar
or tree leaves or Styrofoam may have the same ThOD, their actual oxygen demand may
be substantially different. Further, most wastes are a complex mixture of chemicals (e.g.,
Styrofoam + tree leaves + sugar), present in differing amounts, each with a different level
of biodegradability.
T11 BOD exerted and BOD remaining
Consider the oxidation of organic matter as a function of time. In T11, yt is the CBOD
exerted (oxygen consumed, mg O2/L) and Lt is the CBOD remaining (potential to
consume oxygen, mg O2/L) at any time, t. At t = 0, no CBOD has been exerted (yt=0 = 0)
and all of the potential for oxygen consumption remains (Lt=0 = L0, the ultimate CBOD).
As the oxidation process begins, oxygen is consumed (CBOD is exerted and yt increases)
and the potential to consume oxygen is reduced (CBOD remaining, Lt, decreases). The
rate at which CBOD is exerted is rapid at first, but later slows and eventually approaches
zero as all of the biodegradable organic matter has been oxidized. The total amount of
oxygen consumed in oxidizing the waste is the ultimate CBOD (L0).
What might the ThOD of this sample be? Would it differ for different compounds?
The exponential decline in CBOD remaining (L; potential to consume oxygen) illustrated
in T11 can be modeled as a first-order decay:
dL
 k L L
dt
where kL is the CBOD reaction rate coefficient (day-1). This equation can be integrated to
yield the analytical expression,
Lt  L0e  kLt 
T12 Biodegradability and the nature of kL
The reaction rate coefficient (kL) utilized in CBOD calculations, is a measure of the
biodegradability of a waste. The relationship between kL and the rates of CBOD exertion
(dy/dt) and consumption (dL/dt) is illustrated in T12. Typical ranges for this coefficient
are presented in the table below with reductions in the magnitude of kL as one moves
from untreated sewage → treated sewage → unpolluted river water reflecting progressive
reductions in labile (biodegradable) organic carbon. As with other microbially mediated
processes, values for the CBOD and NBOD reaction rate coefficients vary with
temperature.
Type of sample
Untreated municipal
Treated municipal wastewater
Unpolluted river water
kL (day-1)
0.35 – 0.70
0.10 – 0.35
< 0.05
Returning to the oxygen mass balance, we note that since BOD is expressed in terms of
oxygen consumed (representation by effect), we can quantify the deoxygenation portion
of the dissolved oxygen mass balance using the first order BOD decay equation,
dO2
 kL  L
dt
T13 The DO Mass Balance and the DO sag curve
Consider the effects of L, kL and temperature on the sag curve.
Laboratory determination of BOD
Note that the CBOD exerted at any time t (yt) (T11) is given by the difference between
the ultimate CBOD and the CBOD remaining:
yt  L0  Lt
Substituting Equation 5-31 into Equation 5-32 yields:
yt  L0 1  e  kLt 
and rearranging, yields an expression for the ultimate CBOD:
L0 
yt
1  e 
 kLt
T14 Laboratory determination of BOD
Reoxygenation: the Reaeration Portion of the DO Mass Balance
The driving force for reaeration is the dissolved oxygen deficit, D (mgO2/L), a
representation of how ‘hungry’ the water is for oxygen. The deficit is defined as the
departure from saturation or the difference between the maximum amount of oxygen that
the water can hold and the actual oxygen concentration,
D  DOsat  DOact
The maximum amount of oxygen that the water can hold is termed the saturation
concentration and varies with temperature according to the Henry’s Law,
DO sat  K H  PO2
The value for the Henry’s Law constant for oxygen (KH) decreases as temperature
increases and thus the value of DOsat is lower at higher temperatures.
T15 Oxygen saturation and temperature
Reaeration is a first order reaction based on the reaeration rate coefficient, Ka (d-1), and a
driving force as represented by the dissolved oxygen deficit,
dO2
 ka  D
dt
The magnitude of the reaeration rate coefficient varies with temperature (increases with
increasing temperature) and the turbulence of the stream (increases with increasing
turbulence). Consider the effects of temperature and turbulence on the DO sag curve.
Quantifying the DO Mass Balance
We have now quantified all of the inputs to the DO mass balance,
dO2
 ka  D  k L  L
dt
In practice, the mass balance is written in terms of D because it is really D not DO that
drives the reaeration term,
dD
 k L  L  ka  D
dt
In order to apply this equation in modeling the oxygen resources of a river, the equation
must be integrated,
Dt 
k1  L0
  e k1 t  e k2 t   D0  e k2 t
 k2  k1 
Remembering that protection of the system at the critical point protects the entire system,
T16 The DO Sag Curve and the Critical Point
we would like to know at what “t” does the critical point occur,
tcrit 
 k  D k  k  
1
 ln  2  1  0 2 1  
k

k2  k1
k1  L0

 1 
Applying the Mass Balance
We seek to use the DO mass balance to establish waste treatment objectives that will
maintain a healthy aquatic ecosystem, e.g. achieve a water quality standard for dissolved
oxygen.
T17 Organisms diversity and the DO Sag Curve
T18 Organism abundance and diversity
Let’s examine a case where the mass balance is applied.
T19 Applying the oxygen mass balance
Go To Mass Balance and Kinetics
CMFR SS – 1st Order Decay
3. Lake Water Quality
As with rivers, the management of water quality in lakes requires an understanding of
basic science, a quantitative analysis of perturbation and response (engineering design)
and the application of technologies.
T20 Managing lake water quality
Temperature and the Physical Environment in Lakes
Our starting point for a brief investigation of the science of lakes (limnology) is the
temperature – density relationship for water. Consider what would happen if this were a
linear relationship instead of having a maximum at ~4 °C.
T21 Temperature – density relationship
The density function leads, in lakes, to a phenomenon termed thermal stratification.
Here, the lake divides into a warm, less dense upper layer (the epilimnion or upper lake),
a cold, more dense lower layer (the hypolimnion or lower lake) and a transition region
(the metalimnion or changing lake). Within the metalimnion, the depth where the change
of density (temperature) with depth is greatest is called the thermocline and serves as the
boundary when we seek to represent a lake as a two-layered system.
T22 Thermal stratification
The driving force for this layering effect is most clearly seen in examining vertical
temperature profiles for the spring (isothermal) and summer (stratified) periods.
T23 Temperature – layers and temperature profiles
In temperate climates, lakes stratify in summer and winter (inverse stratification), with
intervening periods of circulation (turnover).
T24 Seasonal cycle of stratification and turnover
T25 Dollar Bay seasonal temperature cycle
Nutrients and Trophic Structure
Start Here
4. Groundwater Quality
What is groundwater
Sources and Types of Groundwater Contamination
For many years, the prevailing view was that contaminants released into the subsurface
would be cleansed by aquifer materials or diluted in the aquifer water. However, a wide
range of toxic substances, including synthetic organic chemicals, trace metals, and
pathogenic microorganisms, have been detected at potentially harmful levels in
groundwater and concern over subsurface contamination has grown.
Groundwater contaminants include:
Pathogens
Nutrients
Inorganic chemicals
Organic chemicals
Radionuclides
from improperly designed or maintained septic tanks
from fertilizers or human waste, e.g. nitrogen and phosphorus
from both natural (arsenic) and anthropogenic (trace metals) sources
from pesticide application, spills, LUST and improper disposal
from the aquifer naturally and nuclear processing facilities
The release of groundwater contaminants may exhibit different areal and temporal
characteristics. For example, pathogens released from a septic tank may originate over a
small area, while pesticide applications may involve > 103 km2. Contaminant releases
may occur over a short time span, such as those associated with a single, essentially
instantaneous chemical spill, or over a longer time span (e.g. decades), such as leakage
from improperly stored chemical wastes. Knowledge of the areal and temporal extent of
a contaminant release is required in the design of remediation plans and preventative
measures.
Pollutant Fate and Transport in Groundwater
Transport in groundwater is mediated by advection and dispersion (diffusion plus
mechanical mixing). Since movement in the subsurface may occur in all three
dimensions simultaneously and can change in time and space, prediction of advective
transport can be quite complicated. However, in a simple, one-dimensional system where
groundwater flow is constant in time and space, the travel time of a contaminant over a
given length can be estimated as:
L
v
where t is the travel time (d), L is length (m) and v (m/d) is the pore (interstitial or
seepage) velocity.
t
The velocity is a function of the driving force for transport (hydraulic gradient, dh/dx,
dimensionless) and the resistance to that transport (hydraulic conductivity, K, m3∙m-2∙d-1).
This relationship,
vDarcy 
Q 
dh 
K  
A 
dx 
is based on the work of the French civil engineer Henri Darcy and is thus termed the
Darcy velocity.
Because the hydraulic conductivity is expressed in terms of the cross-section of the entire
aquifer and groundwater flows only through pores, calculation of the true velocity
requires normalization for porosity (n, dimensionless),
vtrue 
1
dh vDarcy
K 

n
dx
n
Because 0  n  1 , the true velocity will be always be greater than the Darcy velocity.
EXAMPLE. CALCULATION OF GROUNDWATER TRAVEL TIME
A toxic contaminant is released 1 km upstream of a drinking water supply well. The
relevant aquifer properties are: hydraulic conductivity, K = 10-5 m/s; porosity, n = 0.3;
and, hydraulic gradient, dh/dx = -10-2. Determine how long it will take for the
contaminant to reach the drinking water supply well.
Solution
v
 105 m/s 
q
K dh
2
7

 
   10   3.33  10 m/s
n
n dx
0.3


The travel time is then determined from Equation 8-15:
t
1, 000 m
L

 3.00  109 s  95 yr
v 3.33  107 m/s
The long travel time calculated here underscores the fact that groundwater transport can
be slow and may suggest that we do not need to be concerned about the contamination for
a long time period. However, by the time the contamination might be detected at the
well, a large portion of the aquifer and a significant volume of groundwater may be
contaminated as the plume spreads.
Contaminants tend to spread out or mix (disperse) as they are advected with the flowing
groundwater, becoming more dilute and occupying an increasingly larger volume. This
second transport process, dispersion results from the effects of two mechanisms:
diffusion and mechanical mixing.
Diffusion is driven by concentration gradients, while mechanical mixing results from
interactions with the aquifer’s solid matrix (a process analogous to the eddy diffusion
discussed in relation to air and surface water). The contribution of mechanical mixing to
dispersion dominates that of diffusion except at low pore velocities (e.g. in aquifers with
low hydraulic conductivity).
The travel time calculation presented previously assumes that once the contaminant is
released, the contaminant molecules experience identical velocities as they migrate
through the aquifer. In fact, each molecule will be subject to velocities that can range
over many orders of magnitude due to mixing phenomena occurring over a wide range of
length scales. Velocity differences at the pore scale result from friction with the solid
media, an effect manifested to a different degree at different positions over the crosssection. Pore scale velocity differences are also influenced by the flow paths that
molecules travel as they migrate through the aquifer medium. At the aquifer scale,
variations in velocity are caused by differences in hydraulic conductivity resulting in
contaminant arrival at a particular location at different times.
The practical significance of this is that the travel time (for arrival of the front) is shorter
with dispersion than with advection alone. Referring again to the travel time calculation
presented previously, when the effects of dispersion are included, the contaminant
reaches the drinking supply water well at potentially toxic levels orders of magnitude
more quickly than with transport only by advection.
Groundwater contaminants rarely behave conservatively, as they may be degraded by
microorganisms, react chemically or lost as a result of physical processes (e.g.
volatilization, sorption). In many cases, contaminant reactions are beneficial, reducing
contaminant levels. However, in some cases (especially chlorinated solvents), the
reaction products can be at least as toxic as the chemical originally released.
Contaminant interactions with solid materials in the aquifer through sorption are
quantified through a retardation coefficient and serve to slow the rate of chemical
migration.
Groundwater Remediation Strategies
Contaminated groundwater is expensive and time consuming to clean up. This is because:
1) Pollution problems are often discovered years after the initial contamination, offering
the chemical time to impact millions of liters of water and extend kilometers from the
source.
2) The subsurface is physically and chemically heterogeneous. This
heterogeneity leads to slow rates of removal or in-place degradation of contaminants,
even when remediation is aggressive. 3) Little information is usually available
concerning the physical and chemical characteristics of the contaminated aquifer system.
The lack of information, coupled with aquifer heterogeneity and the fact that the
subsurface is essentially invisible, means that the design and implementation of
economical and effective remediation systems presents a challenge.
Groundwater remediation may be accomplished by physical, biological, or chemical
means. Physical remediation usually involves pumping the contaminated groundwater out
of the aquifer, treating it above-ground, and disposing the treated groundwater back to the
impacted aquifer, to a surface water, or to a wastewater collection system. Soil vapor
extraction and air sparging involve pumping and recovering air into and from the vadose
and saturated zones, respectively. Biological remediation consists of injecting electron
acceptors and nutrients. Chemical remediation also involves adding chemicals, to either
enhance abiotic degradation or to immobilize contaminants through chemical
precipitation.