Dispersion of wastes is ignorant of their value.

11.1 Principle 8

Separation may serve different purposes for reuse, recycle, and treatment. For solid waste, food waste (FW) should be separated from other waste because energy could be recovered from the FW. Papers and plastic separated from the solid waste stream could be reused or recycled. For zero water design, it is critical to separate blue (surface and groundwater), green (rainwater), gray (bath and laundry), and black (toilet water). Black water can be further separated into brown (urine) and black (feces). In developing countries, separated urine and feces toilets are promoted by many United Nations programs. N and P could be recovered from the concentrated urine, while anaerobic digestion of feces could generate methane and good organic fertilizer. In developed countries, waterless urinals could save significant amounts of tap water and reduce discharge to the sewer system. As a result, the design criteria of flow rate could be reduced due to water saving by waterless urinals. Separation also holds the key to sludge dewatering and disposal. Water as the final product in WWTP should be separated from any waste produced during the processes such as physicochemical treatment (coagulation) or biological treatment (activated sludge). The former needs to avoid adding chemicals, and the latter needs to separate organic content from water at the front end of a treatment systems. Enhanced separation of organic from wastewater (WW) using a cloth microsieving before WW enters the aeration tank prevents free water from being fixed as cellular water as part of activated sludge. The cloth microsieved solids could be directly diverted to the anaerobic digester to generate more methane to achieve energy‐positive WWTP. As a result, organic loading to the aeration tank would be reduced, and major secondary waste such as activated sludge would also be reduced significantly.

Separation plays important role in reuse, recycling, or treatment. For example, cardboard separated from domestic solid waste (DSW) stream could be used for cardboard reproduction. Fat–oil–grease (FOG) separated from kitchen waste could be codigested with sludge to generate more methane as shown in Table 11.1.

As a treatment strategy, separation technologies are largely dependent on particle size. The particle size is an important factor in the process of water and wastewater treatment (WWT) in the past. Figure 11.1 shows the approximate range of particle size for each specific separation technology.

Figure 11.1 lists sizes for typical particulates in water: viruses, humic acids, clay/silt, bacteria, algae, and sand. Nutrients can be considered to be in the molecules category, with 10⁻¹⁰ m of atomic size. Reverse osmosis can essentially treat all different elements in the figure by allowing water filter through its membrane. One of the major design faults is that tank processes were specifically designed for specific range of particles, coagulation of colloidal particles, and sedimentation for flocculated matters. In WWT, however, different sizes of organic matters ranging from 10⁻³ to 1 mm could be separated in one unit process such as cloth microsieving. Depending upon the opening of the microsieves, 40–60% of organic pollutants could be diverted to anaerobic digester to produce biogas instead of producing sludge. The process would reduce aeration and sludge as secondary waste, and produces more methane for energy‐positive design.

11.2 Challenges and Opportunities

Domestic wastewater (DWW) constitutes 70% of WW in China. With an annual WW volume of 750 million m³, cumulative pollution contributed 225 million tons of COD to Chinese water bodies from 1992 to 2007 (Yu et al., 2015). In rural China, the major pollution sources are DWW and agricultural runoff. From 2001 to 2010, annual discharge of TN was 0.446 and 2.2 million tons and TP 80 000 and 230 000 tons from DWW and the agriculture in China, respectively (Guan et al., 2014). The major challenges faced by rural WWTP are a long construction period, high operation and maintenance (O&M) cost, low technical reliability of a complicated process, high O&M costs, low recycling rates of plants, and geological conditions requiring different designs. More importantly, there are no sustainability design guidelines for rural WWTP and sludge disposal for increasing WW flow rate due to rapid urbanization and population growth.

To solve these urgent WWT challenges, separation is one of the most effective ways to remediate the deterioration of water quality in rural China. There are many opportunities for separation in China. WWT in rural areas of China could be the last big market where only 22% of WW has been treated. China has about 280 000 villages within 1 805 counties, but the WWT rate in rural China is only 13%. In Northern China, groundwater is the major drinking water resource. If WW is not properly treated, contamination of groundwater will spread widely. The 13th Five‐Year Environmental Plan for rural China requires one‐third of villages to have WWT facilities. Currently, 160 million farmers do not have WWT. If $1400 of capital cost for WWT was required for each rural resident, the market capital to build WWT facilities alone would be $240 billion. Indeed, the WWT capital market for rural China could increase exponentially to 3.5, 12, and 20 billion USD in 2016, 2020, and 2025, respectively. In the next decade, the market for WW treatment and the sewer pipe network in China is estimated at 300 billion USD. Separation of gray water from black water would decrease WW at minimal by half. The separated gray water could be used for irrigation of green infrastructures such as wetlands. Separation of urine from feces may offer a way to recover N and P. For example, Ishii and Boyer (2015) used the following unit values in their calculation of urine separation project for a dormitory at the University of Florida (Table 11.2).

According to the assumptions of Table 11.2, for 8 823 population equivalent (PE) annually, the total N and P that can be recovered from 1 920 m³ of urine is 13 200 and 1 020 kg, respectively.

In the past, China designed and built conventional systems such as activated sludge as the major process for towns of more than 1000 people. However, high energy consumption and large amounts of sludge prevent the normal operation of the WWTPs built in recent decades. Two likely extremes exist in rural markets: the technologies were either too low tech and resulted in not enough protection of water resources or too high tech, such as membrane biological reactor (MBR). Unfortunately, a major difficulty also arises due to high capital cost and the intensive energy required to run. As a result, the unit operating cost might be as high as one USD per cubic meter of WW treated. With limited resources of rural WWTP, many of these plants will most likely run into financial difficulty. For these reasons, decentralized and energy positive WWTPs are critical to ensure sustainable WW management in rural China.

In urban China, activated sludge processes have been utilized in more than 95% of the WWTPs in the past four decades. However, most of the WWTPs do not have sludge treatment facility due to the cost of intensive energy. In 2015, cities and counties had a total of 3717 WWTPs with a treatment capacity of about 157 million cubic meters/day. The total treatment capacity of sewage is 48.06 billion cubic meters per year, with 84.1% operating load rate. However, sludge treatment and disposal facilities were only 43.4% at the end of 2013, because sludge treatment and disposal was not mandatory for WWTPs by law. The 12th Five‐Year Plan (FYP) requires the disposal rate of sludge treatment to be 80% by the municipalities directly under the central government, the provincial capitals, and the cities to reach 70% while 30% is required for cities at county or township levels. With consideration of energy footprints and the energy‐positive unit process in WWTP, only 60 of 3717 WWTPs in China have anaerobic digestion facilities of sludge, which is barely 1.7% of the total WWTPs in China. Therefore, it is estimated that sludge treatment and disposal market could be as high as 50 billion USD in the next 10 years. The major sludge disposal problem is the optimization of sludge digestion technology. For example, various composting technologies need to be improved. Sludge heat drying and incineration technology and equipment should be integrated. Land application of treated sludge needs to be monitored. Although incineration, carbon char, and landfill are different options, the land application of anaerobic digested sludge should be the future direction due to the energy production as biogas and pathogen‐free organic fertilizer sludge. According to the US experience, treatment strategy depends upon the PE. To effectively and economically meet discharge standards, three scales are residential, community, and city as shown in Table 11.3, which lists the most suitable treatment technologies and the operation parameters and performance levels required in the United States.

Separation technologies are considered as phase transfer methods and do not destroy organic pollutants. Nevertheless, precipitation, coagulation, flocculation, and sedimentation are still very important unit processes in treating toxic wastewater containing metals. For water reuse, micro-filtration, nano-filtration, and reverse osmosis are critical to ensure the effluent water quality to meet the water reuse standards.

11.3 Precipitation

Chemical precipitation of metal‐contaminated water has long been used in treating WW from industrial applications, such as metal finishing and plating. Effective precipitation is usually accompanied by coagulation, flocculation, and clarification and/or filtration to separate the solid and liquid phases of the waste stream by raising the pH of the solution to lower the metal solubility. Alkali‐precipitating agents such as caustic soda (NaOH), hydrated lime (Ca(OH)₂), and magnesium hydroxide (Mg(OH)₂) are common precipitating agents. The solubility of metal hydroxide precipitates decreases with increasing pH to a minimum value called the isoelectric point. Chemical precipitation depends on several variables:

1. Maintenance of a proper pH range throughout the precipitation reaction and subsequent settling time.
2. Addition of a sufficient excess of treatment ions (precipitating agent) to drive the precipitation reaction to completion.
3. Effective removal of precipitated solids.

Figure 11.2 is a typical schema of a metals treatment system.

After precipitation, coagulation and flocculation are required to destabilize colloidal particle and to grow the particle so that they can settle from WW.

11.4 Coagulation and Flocculation

Coagulation destabilizes colloidal particles, while flocculation refers to growth of destabilized colloidal particles. The stability of colloidal particles is defined by the size, area, and charge of the surface. Colloidal particles attract one another through van der Waals forces and have repulsion forces going against one another due to the zeta potential. When suspended in water, the charge on organic and inorganic colloids is typically negative. Flocculation is the process of bringing together the destabilized or “coagulated” particles to form a larger agglomeration of floc by physical mixing or addition of chemical coagulant aids or both. Electric double layer (EDL) refers to a fixed and a diffusing layer. The potential voltage from the shear of a plane is called the zeta potential. Destabilization reduces the zeta potential. Commonly used coagulants include AlCl₃ (alum), which might cause Alzheimer’s disease due to residual alum. FeCl₃ and FeSO₄ (ferric ion) are a better choice than alum. The rate of flocculation depends upon several factors: (i) the nature of the particle surface, (ii) the presence of charges, (iii) the shape of the particles, and (iv) the density of the particles. Jar tests are used to determine the dose of coagulants through experiment (Figure 11.3).

Mixing provides greater uniformity of the WW feed and disperses precipitating agents, coagulants, and coagulant aids throughout reactor to ensure the most rapid precipitation reactions and subsequent settling of the precipitates. To quantify the degree of mixing, the following factors must be considered: (i) the amount of energy supplied, (ii) the mixing residence time, and (iii) the related turbulence effects of the specific size and shape of the tank. For mechanically stirred mixing basins, G can be calculated as follows:

(11.1)

where

• P = power applied to stirring, W (ft‐lbf/s = HP × 550)
• V = reactor volume m³ (ft³)
• μ = dynamic viscosity N‐s/m² (lbf‐s/ft²)

Power requirements are determined from

(11.2)

where

• P = power applied to stirring, W (ft‐lbf/s)
• C_D = coefficient of drag
• A = paddle area, m² (ft²)
• ρ = fluid density, kg/m³ (slugs/ft³)

For water at 20°C,

where v is the relative velocity of paddles in fluid (m/s [fps]), typically about 0.6–0.75 of paddle tip speed.

11.4.1 Camp–Stein Equation

Velocity gradient provided by the mixer can be estimated by the Camp‐Stein equation:

(11.3)

where

• = power dissipation (mixer power transferred to bulk fluid)
• V = volume of reactor
• μ = dynamic viscosity

After coagulation of the rapid mixing stage, flocculation is based on slow‐to‐enhance contact between particles to promote the growth of destabilized floc. Under the hydraulic force, G, the orthokinetic flocculation colloid number rate decreases according to the following equation:

(11.4)

In‐line static mixers are a type of pipe reactors used for rapid mixing that can be done for any type of flow (laminar, transitional, and turbulent). When comparing both the static and tank mixed reactors, more energy is required in the tank than in the static mixer.

11.4.2 Static and Plug‐flow Reactor Mixers

Figure 11.4 shows the hydraulic characters of PFR, Batch, and CSTR reactors.

11.4.3 Power, Pressure, and Pump in Reactors

Hydraulic power formula is used to calculate the power of each reactor:

(11.5)

where

• γ = specific gravity
• Q = flow rate
• h_L = headloss

Tank reactors form dead zones and short‐circuiting, whereas plug‐flow conditions minimize short‐circuiting and completely avoid dead zones. The US EPA Design Manual (2004) provided good design examples of separation technologies.

After precipitation, sludge handling is an energy‐intensive process. Typically, sludge from water treatment plants is pressure filtrated since it is often difficult to dewater, particularly alum sludge and softening precipitates containing magnesium hydroxide. Gravity‐thickened alum waste is conditioned by the addition of lime slurry. The conditioned sludge is then fed continuously to the pressure filter until filtrate ceases, and the cake is consolidated under high pressure. Filtrate is measured through a weir tank and recycled to the inlet of the water treatment plant. The cake is transported by truck to a disposal site. Alum sludge is conditioned using lime and/or fly ash. Lime dosage is in the range 10–15% of the sludge solids. Ash from an incinerator, or fly ash from a power plant, is applied at a much higher dosage, approximately 100% of dry sludge solids. Polyelectrolytes may also be added to aid coagulation. Fly ash and diatomaceous earth are used for precoating; the latter requires about 50.2 kg/m² of filter area. Under normal operation, cake density is 40–50% solids and has a dense, dry, textured appearance. Alum sludge from surface water treatment is amenable to centrifuge dewatering to about 20% solids. The removal efficiency in a scroll centrifuge ranges from 50 to 95% based on operating conditions and polymer dosage. Alum sludge from surface water coagulation settles to a density in the range of 2–6% solids. Coagulation‐softening mixtures from the treatment of turbid river waters gravity‐thicken approximately as follows: alum–lime sludge, 4–10%; iron–lime settlings, 10–20%; alum–lime wash water, about 4%; and iron–lime backwash, up to 8%. The density achieved in gravity thickening relates to 8%. Lime‐softening precipitates compact more readily than alum floc in a scroll centrifuge. A settled sludge input with 15–25% solids can be dewatered to a solidified cake of 65%. Suspended solids recovery is often 85–90% with polyelectrolyte flocculation. The performance of thickener handling water treatment plant waste varies with the character of the water being treated and the chemicals applied. Calcium carbonate residue from groundwater softening consolidates to 15–25% solids. Solids content and sludge volume are intrinsically linked as in Table 11.4 and Figures 11.9 and 11.10.

Sludge solids content (%)	kg sludge/kg dry solids	m3 sludge/ton dry solids
1	100	100
2	50	50
5	20	20
10	10	9.7
15	6.7	6.4
20	5.0	4.7
30	3.3	3.0
40	2.5	2.2

The regression equation between the log(kg sludge/kg dry solids) and the log(sludge solids content) is

or

The regression equation of log(m³ sludge/ton dry solids) and log(sludge solids content) is

or

11.5 Membrane Filtration Systems

Water membrane filtration systems can be classified as (i) microfiltration, (ii) ultrafiltration, (iii) nanofiltration, (iv) reverse osmosis, and (v) biomembrane reactors. All of these separation technologies are designed to filter a specific range size particle with the corresponding pore size. For example, microfiltration focuses on the largest particle as shown in Figure 11.1. Reverse osmosis is to remove the smallest particles such as ion in the size ranging from several angstrom due to the small membrane pore sizes. It removes cations, acids, and salts at angstroms to nanometers. The filtration systems usually remove the large size particles first. Therefore, microfiltration and ultra‐filtration are typically used for pre‐treatment for nanofiltration or reverse osmosis to prevent fouling.

For water treatment or reclaiming treated wastewater, high pressure is required to force water passing through membrane as permeate and leave concentrate as waste by‐product. Pressure through the pump to aim for permeate performance using hollow fibers is

(11.6)

This is the formula to determine the pressure of the flow.

(11.7)

(11.8)

(11.9)

(11.10)

where: ΔP = transmembrane differential pressure, psi or kPa

dynes/cm²

J = flux, gfd or Lmh

kv = Permeability for overall flow

μ = viscosity constant of water

δ_mem = thickness of the membrane, ft or m

The Long Term 2 Enhanced Surface Water Treatment Rule (LT2ESWTR) requires state to report integrity test of membrane filtration system to earn credit for virus removal (US EPA 2005). The procedure is described in the Membrane Filtration Guidance Manual (US EPA, 2005). Example 11.2 illustrates how to calculate the required technical values to comply with the EPA’s requirements for membrane filtration to earn 3 log virus removal credit. The system is a submerged vacuum-driven hollow-fiber membrane. The upper control limit (UCL) refers to the maximal pressure before the membrane is breached. Log removal value is defined as the log C_effluent – log C_influent.

11.6 Activated Carbon Adsorption

Activated carbon has large adsorption surface area and is easy to design and operate. Therefore, it has been widely used in industrial wastewater treatment. The design is based upon column test results as shown in Example 11.3.

11.7 Anaerobic Membrane Biological Reactor

Bioreactor eliminates BOD and nutrients through either aerobic or anaerobic processes. Anaerobic digestion is a key technology in recovering biogas from sludge. The primary treatment is to remove the suspended solids, the secondary treatment is to degrade BOD, and the tertiary treatment is to remove nutrients. MBR combines the secondary and tertiary treatment together. It could remove both BOD and significant portion of nutrients due to longer solid retention time of large biomass in the MBR. Table 11.9 lists the advantages of membrane bioreactors.

A typical layout of MF within a treatment plant is as follows (Figure 11.13).

According to the different arrangement of membrane bioreactors, they can be classified as submerged MBR (Figure 11.14) versus MBR sidestream (Figure 11.15).

Within the MBR, one of the most popular membranes is the hollow‐fiber MBR (Figure 11.16).

According to loading rate to anaerobic reactors, low and high rates of AD are classified as follows (Table 11.10).

Using the volumetric organic loading rate, which is in kg COD/m³‐day, the volumetric organic loading rate can be calculated as follows:

(11.11)

where

• VOLR = volumetric organic loading rate (kg COD/m³/day)
• S_o = wastewater biodegradable COD (mg/l)
• Q = wastewater flow rate (m³/day)
• V = bioreactor volume (m³)

(11.12)

(11.13)

where

H= reactor height (m)

θ_a = allowable hydraulic retention time (h)

Q = wastewater flow rate (m³/h)

A = surface area of the reactor (m²)

For dilute WW with COD < 1000 mg/l,

The Long Term 2 Enhanced Surface Water Treatment Rule (LT2ESWTR) requires state to report integrate test to earn credit for virus removal (US EPA 2005). The procedure is described in the Membrane Filtration Guidance Manual (US EPA, 2005). Example 11.5 illustrate the calculation on how to calculate the required technical values to earn the 3 log virus removal credit for a submerged vacuum-driven hollow-fiber membrane.

11.8 Air Stripping

Air stripping is a common process used to remove VOCs from water. This is achieved by enabling contact between the liquid (in this case water) and gas (air). Gas is treated to remove the VOCs before it is released into the atmosphere. Air stripping involves the mass transfer of VOCs dissolved in water from the contaminated water to the air phase. This principle is known as Henry’s law:

where, at equilibrium, P_a is the partial pressure of a gas above a liquid and is proportional to the mole fraction of the gas dissolved in the liquid, X_a. Ha is the known Henry’s constant for the gas.

Thermal oxidation and activated carbon are typically used to treat off‐gas that exceeds regulation limits. Thermal oxidation eliminates the contaminant but is a costly, complex process. Low‐profile air strippers can remove 99% of most VOC contaminants. The efficiency of the stripper can be improved, and additional stripping incorporated by increasing the number of trays. Efficiency can also be increased by increasing the airflow. Sieve tray air strippers operate in a similar way to packed‐column air strippers. The difference is that the liquid (water) flows across trays that are perforated with small holes over a weir and through a downcomer to the next lower tray until the treated water flows from the bottom of the stripper. Gas (air) is bubbled through the holes in the trays, stopping the liquid from dripping through them. The VOCs are transferred from the liquid to the gas phase as the air is bubbled through the water on the trays. Figures 11.17 and 11.18 show a simplified diagram of a packed‐column and a low‐profile sieve air stripper.

The theoretical number of trays required is estimated using the following relationship:

where

• X₀ = concentration of contaminant (TCE) in the inlet water phase: 10 mg/l
• X_n = concentration of contaminant (TCE) in the treated water phase: 0.1 mg/l
• N = number of theoretical plates (assumes that the liquid on each plate is completely mixed and where that the vapor leaving the plates is in equilibrium with the liquid)
• H = Henry’s constant (kPa)
• m = slope of equilibrium curve (H/P_t)
• G = kg‐moles air/min
• L = kg‐moles of water/min
• P_t = ambient pressure (kPa)
• S = stripping factor
• Y_n+1 = concentration of volatiles in the air entering the air

For air stripping Y_n+1 = 0 (the concentration of TCE in the air entering the air stripper is zero), and the equation becomes

A packed‐column air stripper consists of a cylindrical column that contains an open‐structured packing material. Water containing the VOCs enters the top of the column and flows down through packing material. At the same time, air flows up through the column (countercurrent flow). As the water and air pass each other, the VOCs are transferred from the water phase to the air phase. The water phase leaves the bottom of the column with most of the VOCs removed. The VOCs that are in the air phase exit from the top of the column. For the packed‐column air stripper, the mass balance equation for VOCs in the liquid versus gas phases is as follows:

(11.14)

where

• L = molar liquid (water) flow per unit of stripper cross‐sectional area
• G = molar gas (air) flow per unit of stripper cross‐sectional area
• x_ai = mole fraction of contaminant in liquid (water) influent
• x_ae = mole fraction of contaminant in liquid (water) effluent
• y_ai = mole fraction of contaminant in gas (air) influent
• y_ae = mole fraction of contaminant in gas (air) effluent

If influent air is not contaminated by VOC, the molar ratio of gas (air) to liquid (water) is

where

• x_ai and L = field measurements and x_ae imposed by ARAR
• p_Te = total pressure of gas (air) effluent (atm)
• P_ae = partial pressure of contaminant a in gas (air) effluent (atm)

From Dalton’s law of partial pressures,

At equilibrium from Henry’s law,

For Pte = 1 atm and 20°C (296.13 K)
Contaminant
Benzene	0.9867	0.2320	4.253 m3/m3
Toluene	0.9000	0.2649	3.397 m3/m3
Trichloroethylene (TCE)	0.8667	0.3797	2.283 m3/m3
Critical contaminant (benzene)

The properties of the liquid (water) at 20°C
	Liquid surface tension
	Liquid viscosity
	Liquid density
	Nominal diameter
	Total surface area
	Critical surface tension for polyethylene packing
	Packing factor

The properties of the gas (air) at 20°C (293.16 K) and 1 atm
	Gas viscosity
	Gas density

11.9 LCA Tools for WWTPs

The core criteria in SEE design are the footprints of EEIS on energy and materials such as water, carbon, and nutrients according to its life cycle of the EEIS. To assess these footprints, many LCA tools have developed as shown in Table 11.18 (Figure 11.21).

Databases that could be used for LCA of WWTP include:

1. E‐PRTR: 28 000 industry pollutants by EU Pollutant Release and Transfer (Yoshida, Water Research, 56, 292–303, 2014).
2. International Reference Life Cycle Data System (ILCD) v2.2 USLCI (Eco‐invent Center, 2007; NREL, 2012).
3. TRACI (Tool for Reduction and Assessment of Chemical and Other Environmental Impacts) (US EPA, 2014).
4. IMPACT World+.

Separation strategy could be effective on different spatial scales because treated WW should be disposed approximate to the reuse. From a sustainability point of view, the energy footprint of WW treatment per cubic meter of WW on different spatial scales such as household (250 GPD), community (0.31 MGD), and city (10.3 MGD) decreases as shown in Figure 11.22 due to economy of scale.

For all the treatment strategies, water reuse is the major positive offset of energy footprint regardless of scale. In conventional AS process, electricity is the major factor contributing to energy consumption. Table 11.19 illustrates the contribution of construction and O&M phases to energy footprint on different scales.

To achieve sustainable design in rural areas, solar energy is the best way to reduce energy footprint by using constructed wetland. A decentralized WWT system utilizing constructed wetlands will be the major technology due to the relatively large residential area for the photo‐remediation or treatment of WW. Solar energy would significantly reduce electricity consumption and treat WW efficiently and economically.

11.10 Capital and O&M Costs of Membrane Filtration

There are three major cost categories: (i) Capital cost includes cost of the system and process reactors. (ii) Operation includes cost of operating the system. (iii) Maintenance covers the cost of fixing and keeping the system up to date. Operation and maintenance costs are generally referred as O&M cost in terms of percentage of capital cost.

11.11 Exercise

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