Energy From Solid And Liquid Wastes - Ix

Lecture No: 23

Anaerobic Sludge Blanket Processes One of the most notable developments in anaerobic treatment process technology was the upflow anaerobic sludge blanket (UASB) reactor in the late 1970s in the Netherlands by Lettinga and his cowworkers ( Lettinga and Vinken, 1980; Lettinga et al; 1980). The principal types of anaerobic sludge blanket processes include (1) the original UASB process and modification of the original design, (2) the anaerobic baffled reactor (ABR) and (3) the anaerobic migrating blanket reactor (AMBR). The ABR process was developed by McCarty and coworkers at Stanford University in the early 1980s (Bachmann et al; 1985). Work in the late 1990s at Iowa State University has led to the development of the AMBR process (Angenent el at; 2000). Of these sludge blanket processes, the UASB process is used most commonly, with over 500 installations treating a wide range of industrial wastewaters. A number of pilot studies have been done with ABR, with a limited number of full-scale installations (Orozco, 1988). In this section all three types of anaerobic sludge blanket processes are described along with their performance, demonstrated process loadings, and key design considerations. The major emphasis is, however, on the UASB process. Upflow Sludge Blanket Reactor Process The basic UASB reactor is illustrated on Fig. 23.1 (a). As shown on Fig. 23.1 (a), influent wastewater is distributed at the bottom of the UASB reactor and travels in an upflow mode through the sludge blanket. Critical elements of UASB reactor design are the influent distribution system, the gas-solids separator, and the effluent withdrawal design. Modifications to the basic UASB design include adding a settling tank (see Fig. 23.1 (b)) or the use of packing material at the top of the reactor (see Fig. 23.1 (c)). Both modifications are intended to provide better solids capture in the system and to prevent the loss of large amounts of UASB reactor solids due to process upsets or changes in the UASB sludge blanket characteristics and density. The use of an external solids capture system to prevent major losses of the system biomass is recommended strongly by Speece (1996). A view of a sludge blanket fixed-film reactor installation is shown on Fig. 23.2.

Fig 23.1. Schematic of the UASB process and some modifications: (a) Original UASB process (b) UASB reactor with sedimentation tank and sludge recycle (c) UASB reactor with internal packing for fixed film attached growth, placed above the sludge blanket

Fig 23.2. View of UASB reactor equipped with internal packing above sludge blanket The key feature of the UASB process that allows the use of high volumetric COD loadings compared to other anaerobic processes is the development of a dense granulated sludge. Because of the granulated sludge floc formation, the solids concentration can range from 50 to 100 g/L at the bottom of the reactor and 5 to 40 g/L in a more diffuse zone at the top of the UASB sludge blanket. The granulated sludge particles have a size rang of 1.0 to 3.0 mm and result in excellent sludge-thickening properties with SVI values less than 20mL/g. several months may be required to develop the granulated sludge, and seed is often supplied

from other facilities to accelerate the system startup. Variation sin morphology were observed for anaerobic granulated sludge developed at 30 and 20°C, but exhibited similar floc size and settling properties (Soto et al; 1997). The development of granulated sludge solids is affected by the wastewater characteristics. Granulation is very successful with high carbohydrate or sugar wastewaters, but less so with wastewaters high in protein, resulting in a more fluffy floc instead (Thaveesri et al; 1994). Other factors affecting the development of granulated solids are pH. Upflow velocity, and nutrient addition (Annachhatre, 1996). The pH should be maintained near 7.0, and a recommended COD:N:P ratio during startup is 300:5:1, while a lower ratio can be used during steady-state operation at 600:5:1. Control of the upflow velocity is recommended during startup by having it high enough to wash out non-flocculent sludge. The presence of other suspended solids in the sludge blanket can also inhibit the density and formation of granulated sludge (Lettinga and Hulshoff Pol, a991). An explanation of the fundamental metabolic conditions associated with granular sludge formation is provided by Speece (1996) based on work by Palns et al. (1987,1990). The explanation is as follows. The formation of dense granulated sludge floc particles is favored under conditions of near neutral pH ,a plug-flow hydraulic regime, a zone of high hydrogen partial pressure, a nonlimiting supply of NH4-N, and a limited amount of the amino acid cysteine. With a high hydrogen concentration and sufficient NH4-N, the bacteria responsible fror granulation may produce other amino acids, but their synthesis is limited by the cysteine supply. Some of the excess amino acids that are produced are thought to be secreted to form extracellular polypeptides which, in turn, will bind organisms together to form the dense pellets or floc granules.

Lecture No: 24 24.1. Anaerobic Sludge Blanket Processes 24.1.1.Up flow Sludge Blanket Reactor Process Design Consideration for UASB Process A comprehensive review of design consideration for UASB reactors has been provided by Lettinga and Hulshoff Pol (1991). Important design considerations are (1) wastewater characteristics in terms of composition and solids content, (2) volumetric organic load, (3) upflow velocity, (4) reactor volume, (5) physical features including the influent distribution system, and (6) gas collection system. 24.1.1.a. Wastewater Characteristics. Wastewaters that contain substances that can adversely affect the sludge granulation, cause foaming, or cause scum formation are of concern. Wastewaters with higher concentrations of proteins and /or fats tend to create more of the above problems. The fraction of particulate versus soluble COD is important in determining the design loadings for UASB reactors as well as determining the applicability of the process. As the fraction of solids in the wastewater increases, the ability to form a dense granulated sludge decreases. At a certain solids concentration (greater than 6 g TSS/L) anaerobic digestion and anaerobic contact processes may be more appropriate. 24.1.1.b. Volumetric Organic Loadings. Typical COD loadings as function of the wastewater strength, fraction of particulate COD in the wastewater, and TSS concentrations in the effluent are summarized in Table 24.1. Removal efficiencies of 90 to 95 percent for COD have been achieved at COD loadings ranging from 12 to 20 kg COD/m3.d on a variety of wastes at 30 to 35°C with UASB reactors. Values for Tfor high-strength wastewater have been as low as 4 to 8 h at these loadings. Where less than 90 percent COD removal and higher-effluent TSS concentrations are acceptable, higher upflow velocities can be used, which will develop a more dense granulated sludge by flushing out other solids. Thus, the higher volumetric COD loadings are shown for this condition.

Table 24.1. Recommended volumetric COD loading for UASB reactors at 30°C to achieve 85 to 95 percent COD removal a

Wastewater COD, mg/L

1000-2000

2000-6000

6000-9000

9000-18000

a

Fraction as particulate COD

Volumetric loading, kg COD/m3.d Flocculent sludge

Granular sludge with high TSS removal

Granular sludge with little TSS Removal

0.10-0.30

2-4

2-4

8-12

0.30-0.60

2-4

2-4

8-14

0.60-1.00

na

na

na

0.10-0.30

3-5

3-5

12-18

0.30-0.60

4-8

2-6

12-24

0.60-1.00

4-8

2-6

na

0.10-0.30

4-6

4-6

15-20

0.30-0.60

5-7

3-7

15-24

0.60-1.00

6-8

3-8

na

0.10-0.30

5-8

4-6

15-24

0.30-0.60

na

3-7

na

0.60-1.00

na

3-7

na

Adapted from Lettinga and Hulshoff Pal (1991)

Note:kg/m3.d x 62.4280 = lb/103 ft3.d Recommended loadings as a function of temperature for wastewaters with mainly soluble COD are presented in Table 24.2. These loadings apply to the sludge blanket volume, and a reactor effectiveness factor of 0.8 to 0.9 as discussed below is used to determine the reactor liquid below the gas collector. The higher loading recommendation for the wastewater containing mainly volatile fatty acids (VFA) is based on the potential of obtaining a more dense granulated sludge. Design T values are also given for treatment of domestic wastewater in Table 24.3 based on pilot-plant experience the

T

value needed is longer than that used in

aerobic processes for secondary treatment for BOD removal. In addition, an aerobic polishing step would likely be needed. The economic benefits of energy savings and lower sludge

production would have to be sufficient to justify the higher capital costs for liquid treatment with a UASB process. Table 24.2. Recommended volumetric organic loadings as a function of temperature for soluble COD substrates for 85 to 95 percent COD removal. Average sludge concentration is 25 g/La

Temperature,°C 15

a

Volumetric loading, kg sCOD/m3. d VFA waste water Non-VFA waste water Range Typical Range Typical 2-4 3 2-3 2

20

4-6

5

2-4

3

25

6-12

6

4-8

4

30

10-18

12

8-12

10

35

15-24

18

12-18

14

40

20-32

25

15-24

18

Adapted from Lettinga and Hulshoff Pal (1991)

Note:kg/m3.d x 62.4280 = lb/103 ft3.d Table 24.3. Applicable hydraulic retention times T for treatment or raw domestic wastewater in a 4-m-high UASB reactora Temperature. °C

a

Average τ, h

Maximum τ, for 4to 6-h peak, h 7-9

16-19

10-14

22-26

7-9

5-7

>26

6-8

4-5

Adapted from Lettinga and Hulshoff Pal (1991)

24.1.1.c. Upflow Velocity. The upflow velocity, based on the flowrate and reactor area, is a critical design parameter. Recommended design velocities are shown in Table 24.4. temporary peak superficial velocities of 6 m/h can be allowed for soluble and partially soluble wastewaters, respectively. For weaker wastewaters the allowable velocity and reactor height will determine the UASB reactor volume, and for stronger wastewaters it will be determined by the volumetric COD loading. The upflow velocity is equal to the feed rate divided by the reactor cross-section area:

Q V=

----- (24.1) A

Where v = design upflow superficial velocity, m/h A = reactor cross-section area, m2 Q = influent flowrate, m3/h Table 24.4. Upflow velocities and reactor heights recommended for UASB reactors a Wastewater type COD nearly 100% soluble

Upflow velocity, m/h Reactor height,m Range Typical Range Typical 10 –3.0 1.5 6-10 8

COD partially soluble

1.0-1.25

1.0

3-7

6

Domestic waste water

0.8-1.0

0.7

3-5

5

a

Adapted from Lettinga and Hulshoff Pal (1991)

Note: m x 3.2808 =ft m/h x 3.2808 = ft/h 24.1.1.d. Reactor Volume and Dimensions. To determine the required reactor volume and dimensions, the organic loading, superficial velocity, and effective treatment volume must be considered. The effective treatment volume is that volume occupied to sludge blanket and active biomass. An additional volume exists between the effective volume and the gas collection unit where some additional solids separations occur and the biomass is dilute. The nominal liquid volume of the reactor based on using a acceptable organic loading is given by. QS 0 Vn =

----- (24.2) L org

Where V n = nominal (effective) liquid volume of reactor, m3 Q = influent flowrate, m3/h So = influent COD, kg COD/m3 L org = organic loading rate, kg COD/m3.d To determine the total liquid volume below the gas collectors, an effectiveness factor is used, which is the fraction occupied by the sludge blanket. Taking into account the

effectiveness factor, which may vary from 0.8 to 0.9, the required total liquid volume of the reactor exclusive of the gas storage area is given by VN V L=

----- (24.3) E

Where V L = total liquid volume of reactor, m 3 V n = nominal liquid volume of reactor, m 3 E = effectiveness factor, unitless Rearranging Eq. (10-13), the area of the reactor is Q A=

----- (24.4) U

The liquid height of the reactor is determined using the following relationship: VL HL=

----- (24.5) A

Where H L = reactor height based on liquid volume, m V L = total liquid reactor volume, m3 A = cross-sectional area, m2 The gas collection volume is in addition to the reactor volume and adds an additional height of 2.5 to 3 m. Thus, the total height of the reactor is H T = H L + HG

----- (24.6)

Where H T = total reactor height, m H L = reactor height based on liquid volume, m H G = reactor height to accommodate gas collection and storage, m 24.1.1.e. Physical Features. The main physical features requiring careful consideration are the feed inlet, gas separation, gas collection, and effluent withdrawal. The inlet and gas separation design are unique to the UASB reactor. The feed in let must be designed to provide uniform distribution and to avoid channeling or the formation of dead zones. The avoidance of channeling is more critical for weaker wastewaters, as there would be less gas production to help mix the sludge

blanket. A number of inlet feed pipes are used to direct flow to different areas of the bottom of the UASB reactor from a common feed source. Access must be provided to clean the pipes in the event of clogging Guidelines for determining the area served by the individual inlet feed pipes as a function of the sludge characteristics and organic loading are provided in Table 24.5. 24.1.1.f. Gas Collection and Solid Separation. The gas solids separator (GSS) is designed to collect the biogas, prevent washout of solids, encourage separation of gas and solid particles, allow for solids to slide back into the sludge blanket zone, and help improve effluent solids removal. A series of upside-down Vshaped baffles is used next to effluent weirs to accomplish the above objectives. Guidelines for the GSS design are summarized in Table 24.6. Table 24.5. Guidelines for sizing the area served by the inlet feed pipes for UASB reactora

Sludge type

COD

Dense flocculent sludge,

Kg/ m 3.d < 1.0

> 40 kg TSS/ m 3

a

loading Area per feed inlet, m 2 0.5–1

1-2

1-2

>2

2-3

Medium flocculent sludge,

<1-2

1-2

20-40 kg TSS/ m 3

>3

2-5

Granular sludge

1-2

0.5-1

2-4

0.5-2

>4

>2

Adapted from Lettinga and Hulshoff Pal (1991)

Note:kg/m3.d x 62.4280 = lb/103 ft3.d Table 24.6. Recommended design considerations for the gas solids separator for UASB reactorsa •

The slope of the settler bottom, i.e, the inclined wall of the gas collector, should be between 45 and 60°

•

The surface area of the apertures between the gas collectors should not be smaller than 15 to 20 percent of the total reactor surface area.

•

The height of the gas collector should be between 1.5 and 2 m at reactor heights of 5-7 m.

•

A liquid-gas interface should be maintained in the gas collector to facilitate the release and collection of gas bubbles and to control scum layer formation

•

The overlap of the baffles installed beneath the apertures should be 100 to 200 mm to avoid upward-flowing gas bubbles entering the settler compartment.

•

Generally scum layer baffles should be installed in front of the effluent weirs.

•

The diameter of the gas exhaust pipes should be sufficient to guarantee the easy removal of the biogas from the gas collection cap, particularly in the case where foaming occurs.

•

In the upper part of the gas cap, antifoam spray nozzles should be installed in the case where the treatment of the wastewater is accompanied by heavy foaming.

Advantages for the UASB process are the high loadings and relatively low detention times possible for anaerobic treatment and the elimination of the cost of packing material. Another major advantage is that the UASB process is, as noted previously, a process are related to those wastewaters that are high in solids content or where their nature prevents the development of the dense granulated sludge. The process design for the UASB process is illustrated in Example 24.1. Example 24.1 UASB Treatment Design. For a UASB treatment process treating an industrial wastewater, determine the (1) size and dimension of the reactor, (2) detention time, (3) reactor SRT; (4) average VSS concentration in biomass zone of the reactor, (5) methane gas production, (6) energy available from methane production, and (7) alkalinity requirements for a wastewater with the characteristics given below to achieve greater than 90 percent soluble COD removal. The wastewater is mainly soluble, containing carbohydrate compounds, and a granular sludge is expected. Assume 50 percent of the influent pCOD and VSS is degraded, 90 percent of the influent sulfate is reduced biologically, and the effluent VSS concentration is 150 g/m3. Assume the design parameters given below and the typical values given in Tables 10-10 and 10-12 are applicable. Wastewater characteristics --------------------------------------------------------Item

Unit

Value

---------------------------------------------------------

Flowrate

m3/d

1000

COD

g/m3

2300

sCOD

g/m3

2000

TSS

g/m3

200

VSS

g/m3

150

Alkalinity

g/m3 as CaCO3 500

SO4

g/m3

Temperature °C

200 30

-----------------------------------------------------Design parameters and assumptions: 1. From Table 10-10, Y = 0.08 g VSS/g COD Kd = 0.03 g VSS/g VSS.d µm = 0.25 g VSS/g VSS.d 2. fd = o.15 g VSS cell debris/g VSS biomass decay 3. Methane production at 35°C = 0.40 L CH4/g COD 4. Reactor volume effectiveness factor = 85 percent 5. Height for gas collection = 2.5 m Solution 1. Determine the reactor volume based on the design organic loading and use of Eq.(1014). a. From Table 24.2 select the average organic loading of 10 k Scod/m3.d Vn = QSo = (1000 m3/d) (2 kg sCOD/m3) (10 kg Scod/m3.d)

Lorg Vn = 200 m3

b. Determine the total reactor liquid volume using Eq.(24.3) VL = Vn = 200m3 = 235m3 E

0.85

2. Determine the reactor dimensions. a. First determine the reactor cross- sectional area using Eq.(24.4) based on

The design superficial velocity. Use the upflow velocity data given in Table 24.4 because the waster is highly soluble, select a velocity of 1.5 m/h. A = Q = ( 1000 m3/d) v

= 27.8 m2

(1.5 m/h) (24 h/d)

A = π D2 = 27.8m2

D = 6m

4 b. Determine the reactor liquid height using Eq. (24.5) HL =VL = 235m3 =8.4m A

27.8m2

c. Determine the total height of the reactor using Eq. (24.6). HT =HL + HG = 8.4m + 2.5m = 10.9m d. Reactor dimensions. Diameter = 6m Height = 10.9m 3. Determine the reactor hydraulic detention time T T= VL = (235m3) (24 h/d) Q

= 5.64 h

(1000 m3/d)

4. Determine the reactor SRT a. The value of the SRT can be estimated by assuming that all the wasted biological solids are in the effluent flow. A conservative consists of biomass. Thus the following relationship applies. QXe = PX,VSS = solids wasted per day Both Q and X, are known. The value of PX,VSS is given by :

PX,VSS = Q (Y) (SO –S) + fd(kd) Q(Y) (SO –S) SRT + Q(nbVSS) - QXe 1 + (kd) SRT b.

1 + (kd) SRT

Develop the data needed to solve the above equation. i The effluent soluble COD concentration at 90% COD removal is S = (1.0 – 0.9) (2000 g/m3 ) = 200 g/m3 ii The effluent nb VSS concentration given that 50 percent of the influent VSS is degraded is:

nbVSS = 0.50(150g/m3) = 75 g/m3 iii The pCOD degraded is pCOD degraded = 0.50(2300 – 2000) g/m3= 150g/m3 iv. Total degradable influent COD, So So = (2000 + 150) g/m3 = 2150 g/m3 Substitute the given parameter values and solve the expression given above for SRT. QXe = (1000m3/d) (150 g/m3). = (1000m3/d) (0.08 g VSS/g COD) [(2150 – 200) g/m3] [ 1 + (0.03 g VSS/g VSS.d) SRT] + (0.15g VSS/g VSS) (0.03g VSS/g VSS.d) (1000m3/d) (0.08 g VSS/g COD) [(2150 – 200 )g/m3]SRT

[ 1 + (0.03 g VSS/g VSS.d) SRT] +(1000 m3/d) (75g/m3)

1500,000 g/d = 156,144,000/[ 1 + (0.03) SRT] + 702,648 SRT/ [ 1 + (0.03)SRT] + 75,000 g/d

SRT = 52d 5. Estimate the effluent soluble COD at an SRT of 52d at 30° using Eq. (7-40) and the given coefficients.

S = KS {1 + (kd) SRT] SRT (Yk – kd) -1

k = µm = (0.25 g VSS/g VSS.d) = 3.125g COD/g VSS.d Y

S=

(0.08g VSS/g COD)

(360mg/L) [1 + 0.03g/g.d)52 d]_________ [(52d) [(0.08g/g) (3.125g/g.d) – (0.03g/g.d)] – 1]

S = 88.3 mg/L

6. Determine if the computed SRT value is adequate.

The fraction of the influent sCOD in effluent = (88.3mg/L) = 0.044 = 4.4%S (2000mg/L)

Because 4.4 percent is less than 10 percent (specified in problem statement), the process SRT is adequate.

7. Determine the average XTSS concentration in biomass zone of the reactor.

a. The value of the XTSScan be estimated by using Eq.(7-35) developed previously in Chap. 7 for the SRT.

SRT = ____V(XTSS)___________________ ( Q – Qw) Xe + QwXR

Because it was assumed that all the wasted solids are in the effluent flow, the Term Qw = 0 and the value of XTSS can be estimated as follows:

SRT ≈ VXTSS

and XTSS ≈ QXe SRT

QXe

V

b. Solve for the value of XTSS, with the volume Vequal to the effective, Volume ,Vn

XTSS ≈ (1000 m3/d) (150g/m3) (52 d) (1 kg/103g) = 39.0 kg/m3 200 m3 The computed value is within the range of solids concentration values given earlier for the UASB process.

8. Determine the methane gas production and energy produced.

a. Determine the COD degraded.

COD = (2150 – 200) g/m3 = 1950 g/m3.

b. Determine the COD removes with sulfate as the electron acceptor. From Sec, 10-3, 0.67 g COD removed/g SO4 reduced

CODSR = 0.09(200 g SO4/m3) (0.67 g COD/g SO4) = 120.6 g/m3

c. Determine the COD used by methanogenic bacteria.

COD MB = (1950 – 120.6) g/m3(1000m3/d) = 1,829,400 g/d

d. Determine the methane production rate.

Methane production at 30°C = (0.04L/g) ( 273.15 + 30) = 0.3935 L/g (273.15 + 35)

Amount of CH4 produced = 0.3935 L/g (1,829,400 g COD /d) = 719,869 L/d = 719.9 m3/d

Total gas volume produced (use 65% methane per Table 10-10) = (719.9m3/d)/(0.65) = 1107.5 m3/d

9.

Determine energy produced from methane. To determine the energy produced, determine the density of methane at 30°C and use the factor 50.1 kJ/g methane ( Table 10-10)

a. Determine density

Density at 35°C = 0.6346 g/L ( Table 10-10)

Methane density at 30°C = (0.6346 g/L) ( 273.15 + 35) = 0.6451 g/L ( 273.15 + 30)

b. Determine energy produced.

Energy produced = (719,869 LCH4/d) (0.6451 g/L) (50.1 kJ/g) = 23.3 x 106 kJ/d

9. Determine the alkalinity requirements. From Table 10-9, the estimated alkalinity concentration required at 30°C and 35% CO2 in the gas phase is 1800 mg/L. Because the alkalinity in the influent is 500 mg/L, the amount of alkalinity that has to be added is

Alkalinity required = (1800 – 500) mg/L as CaCO3 = 1300 mg/L as CaCO3 Daily addition = (1300 g/m3) (1000 m3/d) (1 kg/103g) = 1300 kg/d as CaCO3 Comments: A significant amount of methane gas is produced daily. If the methane can be used for the production of energy by the industrial facility, it could help to offset the cost of adding a considerable amount of alkalinity to maintain the anaerobic reactor pH near 7. Also it is important to note the significance of the effluent solids concentration in determining the system SRT. Anaerobic processes generally produce higher-effluent VSS concentrations compared to aerobic processes. For weak wastewater for which solids production is lower, it may be difficult to maintain long SRT values for high treatment efficiency due to effluent solids loss. Also, in contrast to the situation presented in this example, manual wasting of sludge may be necessary. If the wastewater had a higher-influent COD concentration, and the effluent VSS concentration remained the same, the concentration of solids (X TSS) in the sludge blanket would have to increase and the blanket level would have to be higher. To

avoid a rising sludge blanket, manual wasting of sludge would have to be initiated and SRT value would have to be less than the computed value of 52 d. For design, it is best to assume that the average VSS concentration of the sludge blanket would be less than 25 to 35 kg/m3. 24.1.2 Anaerobic Baffled Reactor In the anaerobic baffled reactor (ABR) process as shown on Fig. 24.1a, baffles are used to direct the flow of wastewater in an upflow mode through a series of sludge blanket rectors. The sludge in the reactor rises and falls with gas production and flow, but moves through the reactor at a slow rate. Various modifications have been made to the ABR to improve performance. The modifications include (1) changes to the baffles design, (2) hybrid reactors where a settler has been used to capture and return solids, or (3) packing has been used in the upper portion of each chamber to capture solids (Barber and Stuckey, 1999). Though granulated sludge is not considered essential for the operation and performance of the ABR process, it has been observed in the ABR process (Boopathy and Tilche, 1992). A number of studies have been done with ABR process at bench and pilot scale for a wide range of wastewaters and temperatures as low as 13°C. An excellent summary of these studies including organic loading rates, temperature, and percent COD removal is provided by Barber and Stuckey, (1999). Typical design loading s for the ABR process are presented in Table 10-17. Many of these studies were operated at T values in the range of 6 to 24 h. the reactor volatile solids concentrations varied form 4 to 20 g/L. Advantages claimed for the ABR process include the following: 1. Simplicity, i.e.; no packing material, no special gas separation method, no moving parts, no mechanical mixing, and little plugging potential. 2. Long SRT possible with low hydraulic retention time T 3. No special biomass characteristic required 4. Wastewaters with a wide variety of constituent characteristics can be treated 5. Staged operation to improve kinetics 6. Stable to shock loads The main limitations at this time (2001) with the ABR process are the lack of the system hydraulics.

Fig 24.1. Schematic views of alternative sludge blanket processes: (a) Anaerobic baffled reactor (ABR) (b) Anaerobic migrating blanket reactor

24.1.3. Anaerobic Migrating Blanket Reactor The anaerobic migrating blanket reactor (AMBR) process is similar to the ABR with the added features of mechanical mixing in each stage and an operating approach to maintain the sludge in the system without resorting to packing or settlers for additional solids capture (see Fig. 24.1 b). Table 24.7 Design and performance results from bench and pilot-scale studies on anaerobic treatment of various wastewaters with the ABR processa

Wastewater

Temperature,

No. of

Influent COD,

COD

Percent

°C

chambers

mg/L

loading

COD

3

kg/m .d

removal

Carbohydrate /protein

35

5

7100-7600

2-10

79-82

Distilling

35

5

51,600

2.2-3.5

90

Carbohydrate /protein

35

5

4000

1-2

94

Molasses

35

3

115,000-900,000

4.3-28

49-88

Swine manure

35

3

58,500

4.0

62-69

Municipal wastewater

18-28

3

264-906

2.2

90

Slaughterhouse

25-30

4

450-550

0.9-4.7

75-90

Pharmaceutical

35

5

20,000

20

36-68

Domestic/industrial

15

8

315

0.9

70

Glucose

35

5

1000-10000

2-20

72-99

a

Adapted from Barber and Stuckey (1999)

Note:kg/m3.d x 62.4280 = lb/103 ft3.d In the AMBR® process, the influent feed point is changed periodically to the effluent side and the effluent withdrawal point is also changed. In this way the sludge blanket remains more uniform in the anaerobic reactor. The flow is reversed when a significant quantity of solids accumulates in the last stage. The AMBR® process has been shown to be feasible from bench-scale testing treating nonfat dry milk (Angenent et al., 2000) at 15 and 20°C. The organic loading rate was varied from 1.0 to 3.0 kg COD/m3,d with hydraulic retention times ranging form 4 to 12 h. At the higher COD loading, the COD removal efficiency was 59 percent at 15°C. At 20°C, COD removals ranged from 80 to 95 percent at COD loading of 1 to 2.0 kg/m3’d. 24.2. Attached Growth Anaerobic Process Upflow attached growth anaerobic treatment reactors differ by the type of packing used and the degree of bed expansion. Three types of upflow attached growth processes are illustrated on Fig. 24.2. In the upflow packed-bed reactor (see Fig. 24.2 a) the packing is fixed and the wastewater flows up through the interstitial spaces between the packing and biogrowth. Effluent recycle is generally not used for the packed-bed reactor except for highstrength wastewaters. While the first upflow anaerobic packed-bed processes contained rock, a variety of designs employing synthetic plastic packing are used currently. The anaerobic expanded-bed reactor (see Fig. 24.2 b) uses a fine-grain sand to support biofilm growth. Recycle is used to provide upflow velocities, resulting in 20 percent bed expansion. Higher upflow velocities are used for fluidized-bed- anaerobic reactors (see Fig. 24.2 c), which also contain a fine-grain packing. In fluidized-bed systems, both fluidization and mixing of the packing material occurs. The expanded and fluidized-bed reactors have more surface area per reactor volume for biomass growth and better mass transfer than the upflow packed-bed reactor, but have lower solids capture. These different systems are described in more detail in the following paragraphs, along with their typical design conditions and COD loadings.

Fig 24.2. Upflow anaerobic attached growth treatment reactors: (a) Anaerobic upflow packed bed reactor (b) Anaerobic expanded bed reactor (c) Anaerobic fluidized bed reactor 24.2.1. Upflow Packed-Bed Attached Growth Reactor Full-scale upflow packed-bed anaerobic filters are used in cylindrical or rectangular tanks at widths and diameters ranging form 2 to 8 m and heights form 3 to 13 m (see Fig. 24.2 a). Packing material placement may be in the entire depth or, for hybrid designs, only in the upper 50 to 70 percent. The most common packing materials are corrugated plastic crossflow or tubular modules and plastic pall rings. The specific surface area of the packing averages about 100 m2/m3 and, based on research results, no performance improvements were observed at higher packing densities (Song and Young, 1986). Cross –flow packing appeared to provide higher process performance efficiency over randomly packed medium (Young and Yang, 1989). Typical COD loadings used, hydraulic retention times, and COD removal efficiencies for upflow packed-bed anaerobic reactors are reported in Table 24.8. At loadings of 1.0 to 6.0 kg COD/m3.d, process efficiencies up to 90 percent are shown for high-strength wastewaters. A large portion of the biomass responsible for treatment in the upflow attached growth anaerobic processes is loosely held in the packing void spaces and not just attached to the packing material (Young and Dehab, 1983). Low upflow velocities are generally used to prevent washing out the biomass. Over time, solids and biomass will accumulate in the packing to cause plugging and flow short circuiting. At this point, solids must be removed by flushing and draining the packing.

Advantages of upflow attached growth anaerobic reactors are high COD loadings, relatively small reactor volumes, and operational simplicity. The main limitations are the cost of the packing material and operational simplicity. The main limitations are the cost of the packing material and operational problems and maintenance associated with solids accumulation and possible packing plugging. The process is best suited for wastewaters with low suspended solids concentrations. Table 24.8. Examples of process operating conditions and performance for upflow attached growth anaerobic reactorsa Wastewater

a

Packing

Temp,

COD

type

°C

loading

Recycle

COD

ratio, R/Q

remove

1.2

5.0

d, % 61

τ, d

Guar gum

Pall rings

37

kg/m3.d 7.7

Chemical Processing

Pall rings

37

12-15

0.9-1.3

5.0

80-90

Pall rings

15-25

0.1-1.2

0.5- 0.75

0

50-70

Domestic

Tubular

37

0.2-0.7

25-37

0

90-96

Landfill Leachate

Cross- flow

35

1.5-2.5

2.0-3.0

0.25

89

Food canning

Cross- flow

30

4-6

1.8-2.5

0

90

Soft drink

2-stage

Adapted from Young (1991)

Note:kg/m3.d x 62.4280 = lb/103 ft3.d 24.2.2. Upflow Attached Growth Anaerobic Expanded-Bed Reactor In the upflow attached growth anaerobic expanded-bed reactor (AEBR) process (see Fig. 24.2 b), the packing material is generally silica sand with a diameter in the range of 0.2 to 0.5 mm and specific gravity of 2.65. For operation with about 20 percent bed expansion, an upflow velocity of about 2 m/h is used. The smaller packing provides a greater surface area per unit volume, theoretically supporting a greater amount of biomass growth. The packing void fraction is about 50 percent when expanded and the specific surface area is in the range of 10,000 m2/m3 (Malina and Pohland, 1992). With such a small packing and void volume, the expanded-bed operation is necessary to prevent plugging. Because the expanded-bed system is not fully fluidized, some solids are trapped and some degree of solids degradation occurs (Morris and Jewell, 1981). Most application for the AEBR treatment process have been for the treatment of domestic wastewater (Jewell, 1987). More recently, Collins et al. (1998)’ have investigated

the application of the AEBR for treaqting domestic wastewater at low temperatures. The results of these studies are summarized in Table 24.9, where the date for COD loading and removal efficiency at 20°C are within the same order of magnitude as shown for upflow packed anaerobic bed reactors treating industrial wastewater. The advantages and limitations are similar to those for the fluidized-bed reactor discussed next. Table 24.9. Performance of bench-scale upflow anaerobic expanded –bed reactor treating domestic wastewater at low temperaturea

a

Temperature, °C

Organic loading,

Percent COD removal

20

kg COD/m3.d 4.4

89

15

4.0

80

10

0.4

71

5

0.3

35

Adapted from Alderman et al. (1998)

Note:kg/m3.d x 62.4280 = lb/103 ft3.d 24.2.3. Attached Growth Anaerobic Fluidized-Bed Reactor The attached growth anaerobic fluidized-bed reactor (FBR) (see Fig. 24.2c) is similar in physical design to the upflow expanded-bed reactor. The packing size (~0.3 mm sand) is similar to the expanded-bed reactor, but the FBR is operated at higher upflow liquid velocities of about 20 m/h to provide about 100 percent bed expansion. Effluent recycle is ised to provide sufficient upflow velocity. Reactor depth ranges from 4 to 6m. Besides sand, other packing materials have been considered for use in FBRS including diatomaceous earth, anion and cation exchange resins, and activated carbon (Wang et al., Kindzierski et al., 1992). Though a greater biomass concentration could be developed with more porous diatomaceous earth packing, better performance did no result when compared to sand. Diffusion limitations into the biomass within the porous packing structure may explain the lack of improved treatment efficiency. Activated carbon has been used in many anaerobic FBRS for treating industrial and hazardous waste streams (Hickey et al., 1991; Iza, 1991). The mean diameter of the granular activated carbon particles is 0.6 to 0.8 mm and upflow velocities of 20 to 24 m/h are used. Many benefits associated with using activated carbon in the anaerobic FBR (Wang et al., 1986; Fox et al., 1998) are listed in Table 24.10. The main limitation with activated carbon is

the higher cost, but for certain types of industrial and hazardous waste streams the use of activated carbon is a necessity. Table 24.10 Benefits of using granular activated carbon (GAC) versus sand as the packing material in anaerobic FBRS -----------------------------------------------------------------------------------------------------•

Higher biomass concentration maintained due to porous structure of GAC

•

Adsorption properties help prevent toxic and inhibitory substances from decreasing biological treatment performance. Concentration of degradable substrate can be reduced to below toxic levels Other toxic substances can be removed to protect bacteria

•

Adsorption properties may minimize shock loads by sorption of increased organics

•

Adsorption properties may help acclimate and enhance biomass degradation of toxic compounds by providing more time of exposure

--------------------------------------------------------------------------------------------------Solids capture is minimal in the anaerobic FBR due to the high turbulence and thin biofilms developed. With little solids capture, the process is better suited for wastewaters with mainly soluble COD. Solids discharged in the effluent from sloughed biofilm are minimized by controlling the biofilm inventory in the reactor. As biomass accumulates on the FBR packing, the net particle density decreases and the particle migrates to the top of the reactor. Periodic removal of these solids can control biofilm sloughing and minimize effluent TSS concentrations. The removed particles are mechanically processed to separate biomass from the sand, which is returned to the FBR. Startup of anaerobic FBR S must be done with more care than the other types of highrate anaerobic reactors. A higher hydraulic application rate is recommended at first to select for bacteria that more readily attach to the reactor packing under the highly turbulent conditions (Sutton and Huss, 1984; Denac and Dunn, 1998). The startup time can take 3 to 6 months. In a laboratory study by Tay and Zhang (2000) the startup and performance of an aerobic FBR was compared to that for an anaerobic UASB reactor and upflow packed-bed reactor. All three could achieve COD loading of 10 kg COD/m3.d at 35° C in 3 months with an influent COD concentration of 5000 mg/L (primarily glucose) and a T of 12 h. The COD removal efficiency was best for the FBR and UASB reactors, about 96 percent compared to 90 percent for the upflow packed bed reactor.

Process COD loading values of 10 to 20 kg COD/m3.d are feasible for anaerobic FBRS with greater than 90 percent COD removal, depending on the type of wastewater. Treatment performance is higher for FBRS than upflow packed-bed reactors at higher loadings due to a greater mass transfer rate due to the turbulent mixing. Reactor biomass concentrations of 15 to 20 g/L can be established in anaerobic FBRS (Malina and Pohland, 1992). Anaerobic FBR loadings and performance data are presented in Table 24.11. Benchscale or pilot-plant studies are normally done before establishing full-scale design loadings. Table 24.11. Examples of process operating conditions and performance for anaerobic FBRSa τ, d

COD

kg/m3.d 42

24

removed, % 70

35

8.2

105

99

Milk

37

3-5

18-12

71-85

Molasses

36

12-30

3-8

50-95

Glucose

35

10

12

95

Sulfite, pulp

35

3-18

3-62

60-80

Wastewater

a

Temperature,

COD loading

Citric acid

°C 35

Starch, whey

Adapted from Denac and Dunn (1998).

Note:kg/m3.d x 62.4280 = lb/103 ft3.d The advantages for the anaerobic FBR process include the ability to provide high biomass concentrations and relatively high organic loadings, high mass transfer characteristics, the ability to handle shock loads due to its mixing and dilution with recycle, and minimal space requirements. The process is best suited for soluble wastewaters due to its inability to capture solids. Care must also be taken in the inlet and outlet designs to assure good flow distribution. Other disadvantages include the pumping power required to operate the fluidized bed, the cost of reactor packing, the need to control the packing level and wasting biogrowth, and the length of startup time. 24.2.4. Downflow Attached Growth Processes Downflow attached growth anaerobic processes, as illustrated on Fig. 24.3, have been applied for treatment of high-strength wastewaters using a variety of packing materials including cinder block, random plastic, and tubular plastic. Packing heights are in the range of 2 to 4 m, and systems are designed to allow recirculation of the reactor effluent.

Fig 24.3. Downflow attached growth anaerobic treatment reactor with plastic packing

Anaerobic downflow attached growth reactors can be operated at loadings in the range of 5 to 10 kg COD/m3.d for easily degradable wastewaters. Operating performance varies for different wastewaters as show in Table 24.12. Because plugging of the packing can be of concern, the use of a packing material with a high void volume, such as the vertical plastic packing used in tower trickling filters, is recommended.

Table 24.12. Examples of process operating conditions and performance for anaerobic downflow attached growth applicationsa Wastewater

Temp,

COD loading

τ, d

COD

Citrus

°C 38

kg/m3.d 1-6

24-144

removed, % 40-80

Cheese whey

35

5-22

2-8

92-97

Sludge heat-treatment liquor

40

20-30

Brewery

35

20

1-2

76

Molasses

35

2-13

14-112

56-80

Piggery slurry

35

5-25

0.9-6.0

40-60

58

°Adapted in part from Speece (1996); Lomas et al.(1999); Fredericks et al., (1994); and Jhung and Choi (1995). Note:kg/m3.d x 62.4280 = lb/103 ft3.d The major advantages for the downflow attached growth process, where a higher void space packing material is used, are a simpler inlet flow distribution design, no plugging

problems, and a simple operation. For systems with aerobic treatment following anaerobic treatment, the solids are captured in the aerobic process and thus do no accumulate in the attached growth process. Similar to the other anaerobic processes used for high-strength industrial wastewater treatment, benefits include the ability to treat high COD loadings with relatively small reactor volume sizes. Disadvantages include the cost of the packing material, and the somewhat lower organic loading rates to achieve the same treatment efficiency as the UASB and FBR processes.

Lecture No: 25

Introduction to Risk Assessment When health effects can occur as the result of an environmental action, risk analysis is used to quantify the corresponding risks. Typically, a complete risk analysis is divided into two parts: (1) risk assessment and (2) risk management. Risk assessment involves the study and analysis of the potential effect of certain hazards to human health. Using statistical information on cause and effect, risk assessment is intended to be a tool for making informed decisions. Risk management is the process of reducing risks that are determined to be unacceptable. Both risk assessment and risk management and risk management are introduced in the following discussion to provide a perspective on the approaches used in risk analysis. Noncarcinogenic effects and ecological risk assessment are also considered briefly. Additional details may be found in Pepper et al. (1996) and U.S. EPA (1990). Finally, risk assessment for water reuse is discussed with respect to enteric viruses, which are a major concern in industrialized countries. Risk Assessment Environmental risk analysis takes place in four discrete steps, as diagramed on Fig. 25.1. (Neely, 1994). 1. Hazard identification 2. Exposure assessment 3. Dose response assessment 4. Risk characterization These four steps are described below, along with a brief discussion of the use of risk assessment in setting standards. Hazard Identification. This step involves weighing the available evidence and determining whether a substance or constituent exhibits a particular adverse health hazard. As part of hazard identification, evidence is gathered on the potential for a substance to cause adverse health effects in humans or unacceptable environmental impacts. For humans, the principal sources for this information are clinical studies, controlled epidemiological studies, experimental animal studies, and from evidence gathered from accidents and natural disasters.

Exposure Assessment. Exposure is the process by which an organism comes into contact with a hazard; exposure or access is what bridges the gap between a hazard and a risk (Kolluru et al., 1996). For humans, exposure can occur through different pathways including inhalation of air, ingestion of water or food, absorption through the skin via dermal contact, or absorption through the skin via radiation. The key steps in exposure assessment are identification of a potential receptor population, evaluation of exposure pathways and routes, and quantification of exposures. For example, an exposure scenario to assess the impacts of drinking groundwater that contains a known amount of trichloroethylene (TCE) would be as follows: an adult weighing 70 kg drinks 2L of groundwater containing 50 µg/L of TCE every day for 70 years. In this case, the adult is the receptor, drinking groundwater is the pathway, and drinking 2.0 L/d containing 50 µg/L of TCE for 70 years is the quantification of exposure.

Fig 25.1. Definition sketch for the conduct of health effects risk assessment Dose-Response Assessment. The fundamental goal of a dose-response assessment is to define a relationship (typically mathematical) between he amount of a toxic constituent to which a human is exposed and the risk that there will be an unhealthy response to that dose in humans. The relative sensitivity of epidemiological studies in defining excess risk is illustrated on Fig. 25.2.a. Typical does-response relationships for carcinogenic and noncarcinogenic constituent are illustrated on Fig. 25.2.b. It should be noted that it is assumed that there is no threshold for potentially carcinogenic constituents. Although the does-response curve for a carcinogenic constituent is shown passing through the origin, data

are not available at extremely low doses (see Fig. 25.3.a). Therefore, mathematical models have been developed to define the dose response at low concentrations. Typical doseresponse models that have been proposed and used for human exposure include (1) the single-hit model, (2) the multistage model, (3) the linear multistage model, (4) the multihit model, and (5) the probit model. The characteristics of these models are summarized in Table 25.1.

Fig 25.2. Definition sketches for risk assessment: (a) Relative sensitivity of epidemiological studies in defining excess risk (b) Dose-response curves for carcinogenic and non-carcinogenic constituents

Fig 25.3. Definition sketch for dose- response curves: (a) Illustration of where data are available and where data are required. (b) Two different models used to define the dose-response relationship

The mathematical function used to describe the relationship between risk and dose for the single-hit mode is P(d) = 1 – exp [- (q0 – q1d ) ---------( 25.1) Where P (d) = lifetime risk (probability) of developing cancer q0 and q1 = empirical parameters picked to fit the data d = dose The mathematical formulation used to describe the relationship between risk and dose for the multistage model is n

P (d) = 1 - exp [ - Σ qi di] ---------( 25.2) i =0

where P (d) = lifetime risk (probability) of developing cancer qi = positive empirical parameters picked to fit the data d = dose The relationship between these two models and the risk data given on Fig.25.3.a is shown on Fig.25.3.b. A variety of models have also been proposed to define the risk associated with low levels of microbial pathogens (Haas, 1983). The U.S.EPA has defined lifetime risk as follows: Lifetime risk = CDI x PF

---------(25.3)

Where CDI = chronic daily intake over a 70-year lifetime, mg/kg-d PF = potency factor, (mg/kg.d)-1 The chronic daily intake (CDI) is computed as follows: Total does, mg CDI = --------------------------------------- ---------(25.4) (Body weight, kg) (Lifetime, d)

In its most general form, the total dose is defined as constituent Total dose =

intake x

concentration

exposure x

rate

absorption x

duration

---------(25.5) factor

Table : 25.1. Models used to assess nonthreshold effects of toxic constituents a Model b One-hit Multistage

Description A single exposure can lead to the development of a tumor The formation of a tumor is the result of a sequence of biological

Linear multistage

events Modification of the multistage model. The model is linear at low doses with a constant of proportionality that statistically will

Multihit Probit

produce less than 5 percent chance of underestimating risk Several interactions are required before cell becomes transformed Tolerance of exposed population is assumed to follow a longnormal (probit)distribution

a

Adapted from Cockerham and Shane (1994), Pepper et al. (1996).

b

In all of the models cited above, it is assumed that exposure to the toxic constituent will

always produce an effect regardless of the dose. Recommended standard values for daily intake calculations have also been developed by EPA. The average body weights used for an adult and child are 70 and 10 kg, respectively, and the corresponding rates of water ingestion are 2 and 1 liter(s) per day (U.S. EPA, 1986b). The potency factor PF, often identified as the slope factor, is the slope of the doseresponse curve, at very low doses (see Fig. 25.2.b). The U.S. EOA has selected the linear multistage model as the basis for assessing risk. In effect, the PF corresponds to the risk resulting from a lifetime average dose of 1.0 mg/kg.d. The U.S. EPA maintains a database on toxic substances known as the Integrated Risk Information System (IRIS) (U.S. EPA, 1996). Typical toxicity data for several chemical constituents are reported in Table 25.2.

Table: 25.2 Toxicity data for selected potential carcinogenic chemical constituents a, b

Chemical constituent CASRN

c

Potency factor, PF Oral route,

Inhalation

(µg/kg.d)-1 3.0 E-2 1.54 to 5.45 E-5 Na 1.6 E-4 3.2 E-2 9.1 E-3 3.0 E-1

route,

Arsenic, inorganic Benzene Bromate Chloroform Dieldrin Heptachlo N-

7440-38-2 71-43-2 15541-45-4 67-66-3 60-57-1 76-44-8 55-18-5

(mg/kg.d)-1 1.5 E+0 1.5 to 5.5 E-2 7 E-1 6.1 E-3 1.6 E+1 4.5 E+0 1.2 E+2

Nitrosodienthylamin N-

62-75-9

5.1 E+1

9.8 E-2

Nitrosodimethlamine Vinyl chlorided 75-01-4

7.2 E-1

3.1 to 6.2 E-5

a

U.S. EPA IRIS database (2996) (http://www.epa.gov/iris).

b

Because the data in the IRIS database is being revised continuously, it is important to check the database for the most current values.

c

Chemical Abstracts Service Registry Number.

d

Continuous lifetime exposure during adulthood. Comparing the magnitude of the given values listed in Table 25.2, the relative potency

of the chemical constituents can be assessed (e.g., for the oral route, the potency of arsenic is about 245 times that of chloroform). Because of the numerous uncertainties involved in the development of the database, it is important to remember that the values given in the IRIS database cannot be used to predict the incidence of human disease or the type of effects a given chemical constituent will have on an individual The data included extrapolations of animal data to humans and from high experimental dosages to the low environmental dosages encountered in real life. Use of the data given in Table 25.2 is illustrated in Example 25.1. Example 25.1 Risk Assessment for Drinking Groundwater Containing Trace Amounts of NNitrosodimethylamine (NDMA), Estimate the incremental cancer risk for an adult associated with drinking 2 L per day of groundwater containing 2.0 µg/L of N-Nitrosodimethlyamine (NDMA) using data from the Integrated Risk Information System (IRIS) (U.S.EPA, 1996).

To limit NDMA exposure to acceptable cancer risk of 1 in 1,00,000 determine the concentration of NDMA that can be allowed in extracted groundwater. Solution 1.

Compute the CDI using Eq. 25.4.

Average daily dose, mg/d CDI = ------------------------------Body weight, kg (2.0 µg/L) (2L/d) (1 mg/103 µg) CDI = -------------------------------------- = 0.57 x 10 –4 mg/ kg.d 70 kg

2.

Compute the lifetime risk using Eq. 13-3 and date form Table 25.2.

Incremental lifetime risk = CDI x PF The potency factor from Table 25.2 for the oral route for NDMA is 5.1 x 10 (mg/kg.d)-1 Thus, Incremental lifetime risk = (0.57 x 10 –4 mg/kg.d) [5.1 x 10 (mg/kg.d)-1] = 2.9 x 10-3 From the results of this analysis, the estimated probability of developing elevated cancer risks as a result of drinking the groundwater containing 2.0 µg/L of NDMA is 2.9 per 1000 persons. 3.

Determine the concentration of NDMA to limit the acceptable cancer risk to 1

in 1,00,000 a.

Estimate the CDI

10-5 = (CDI) [5.1 x 10 (mg/kg.d)-1] CDI = 1.96 x 10-7 mg/kg.d b.

Estimate the concentration of NDMA

(Cµg/L) (2L/d) (1 mg/103 µg) ----------------------------------- = 1.96 x 10-7 mg/kg.d 70 kg

C = 0.0069 µg/L = 6.9 ng/L

Comment: The U.S. EPA-proposed drinking water standard for NDMA is 2.0 ng/L. Because of concern over the carcinogenocity of NDMA, the Ontario, Canada, Drinking Water Objective has been set at 9 ng/L, based on a cancer risk to 1 in 1,00,000 estimated by the U.S.EPA (Andrews and Taguchi, 2001). In addition to the carcinogenic dose-response information, the U.S. EPA has developed reference doses (RfD) for a number of constituents based on the assumption that thresholds exist for certain toxic effects (see Fig. 25.2), such as cellular necrosis (localized death of living tissue), but may not exist for other toxic effects, such as carcinogenicity. IN general, RfDs are established based on reported results from human epidemiological data, long-term animal studies, and other available toxicological information. The RfD values represent an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human population (including sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects during a lifetime (U.S. EPA, 1989a). RfD values are available in the IRIS database (U.S. EPA, 1996) and in the Health Effects Assessment Summary Tables (U.S. EPA, 1991). The RfD is used a sa reference point for assessing the potential effects of other doses. Usually, doses that are less than the RfD are not likely to be associated with health risks. As the frequency of exposure exceeds the RfD and the size of excess increases, the probability increases that adverse health effects may be observed in a human population. The RfD is derived using the following formula: NOAEL or LOAEL RfD = --------------------------(UF1x UF2…) x MF

Where NOAEL

=

no observable adverse effect level

LOAEL

=

lowest observable adverse effect level

UF1, UF2

=

uncertainty factors

MF

=

modifying factor

In the above equation, uncertainty factors are based on experimental species, effects, and duration of the study, while modifying factor represent professional assessments

reflecting the confidence in the study. The LOAEL is used only when a suitable NOAWL is unavailable. It must be recognized that the present state of knowledge concerning the impacts of specific constituents is incomplete. Thus, each step in risk assessment involves uncertainty. In hazard identification, most assessments depend on animal test and yet animal biological systems are different form human ones. In dose response, it is often unknown whether safe levels or thresholds exist for any toxic chemical. Exposure assessment usually involves modeling, with the attendant uncertainty as to substance release, release characteristics, meteorology, and hydrology. Because of the uncertainties associated with any risk assessment, the results of such an analysis should only be used as a guide in decision making (Haas et al., 1999). Risk Characterization The final step in risk assessment is risk characterization, in which the question of who is affected and what are the likely effects are defined to the extent they are known. Risk characterization involves the integration of exposure and does-response assessments to arrive at the quantitative probabilities that effects will occur in humans for a given set of exposure conditions. Risk Assessment in Standard Setting Examples of risk assessment in wastewater management include health effects from consuming highly treated reclaimed water and health and environmental effects from land application of biosolids. An acceptable risk of 1 in 10,000 is often used in environmental risk assessment (U.S. EPA, 1986a). Risks of less than 1 in 10,000 are considered minimal. Risk Management Risk management involves the development of standards and guidelines and management strategies for specific constituents including both toxic constituents and infectious agents. For example, if a toxic constituent or infectious microorganisms are present at higher than the maximum allowable concentration based on the risk assessment, risk management involves the determination of what management and/or

technology is necessary to limit the risk to an acceptable level. Thus, the development and screening of alternatives; selection, design, and implementation; and monitoring and review are important elements of health risk management. Ecological Risk Assessment Ecological risk assessment is similar to risk assessment for humans in that the ecological effects of exposure to one or more stressors are assessed. A stressor is defined as a substance, circumstance, or energy field that can cause an adverse effect on a biological system. It should be noted that ecological risk assessments are undertaken for a variety of reasons such as to assess the potential impacts of the discharge of treated effluent to an existing wetland or applying biosolids on land. The framework for ecological risk assessment is illustrated on Fig. 25.4 involving (1) problem formulation, in which the characteristics of the stressor are identified, (2) identification and characterization of the ecosystem at risk and the exposure modes, (3) identification of likely ecological risks, and (4) risk characterization, in which all of the information and data are integrated along with input from the risk manager (U.S.EPA, 1992b). Because the field of ecological risk assessment is continually undergoing change, the latest reports and publications should be consulted. Risk Assessment for Water Reuse In less-developed countries where advanced levels of wastewater treatment are not possible or are economically out of reach, a number of investigators have sought to assess the risk of using reclaimed water of varying quality different reuse applications by controlling possible transmission routes of excreta-related infections. The WHO Health Guidelines for the Use of Wastewater in Agriculture and Aquaculture (1989) is such an example where high concentrations of pathogens exist in wastewater and partially treated effluents.

Figure 25.4. Definition sketch for the conduct of ecological risk assessment

In the United States, the constituents in reclaimed water that have received the most attention are enteric viruses because of their low-dose infectivity, long-term survival in the environment, difficulties in monitoring them, and their low removal and inactivation efficacy in conventional wastewater treatment. However, more recently, other inorganic and organic constituents that may be present in treated wastewater such as arsenic and NDMA have received considerable attention because of their significant potency factors (see Table 25.2). Health risks associated with enteric viruses in reclaimed water that are typically encountered in the California water reuse conditions were analyzed by a quantitative microbial risk assessment approach (Tanaka et al., 1998). Past monitoring data from four wastewater treatment facilities in California on enteric virus concentrations in unchlorinated secondary effluents were used. To assess potential health risks associated with the use of reclaimed water in various reuse applications, four exposure scenarios were tested: (1) golf course irrigation, (2) food crop irrigation, (3) recreational impoundments, and (4) groundwater recharge. Because enteric virus concentrations in unchlorinated secondary

effluents were found to vary over a wide range, characterizing their variability was found to be extremely important in this study. Two concepts related to safety of water reuse were used: (1) the reliability, defined as the probability that the risk of infection does not exceed an acceptable risk, and (2) the expectation, defined by acceptable annual risk in which exposure to the enteric viruses may be estimated stochastically by numerical simulation such as the Monte Carlo methods. In the U.S. EPA Surface Water Treatment Rule (SWTR) (U.S. EPA, 1989b), it is assumed that one infection per 10,000 population per year due to pathogens is acceptable in the public water supply. Therefore, if 10 – 4 annual risk of infection (less than or equal to 1 infection per 10,000 population per year) is set as an acceptable risk for water reclamation and reuse, the reliability (percent of time that infection risk due to exposure to enteric viruses in reclaimed water is less than the acceptable risk) is presented in Table 25.3 From the results of the analysis presented in Table 25.3, the reliability or relative safety of water reuse can be assessed in comparison to the domestic water supply meeting the SWTR. When the effluent from the full treatment with high chlorine does of about 10 mg/L is used, there is virtually no difference in terms of a probability of enteric virus infection whether reclaimed water or domestic water is used for golf course irrigation, crop irrigation, and groundwater recharge. However, depending on the water quality of the secondary effluent, there is a considerable difference in water reuse for recreational impoundment where body contact sports and swimming may take plant. Similar observations can be made for the use of chlorinated secondary effluent and the reclaimed water from contact filtration with low chlorine does of less than 5 mg/L. The utility of the study conducted by Tanaka et al. (1998), as described above, was that their findings were able to provide a basis for conservative assessment of microbiological requirements promulgated in the California Wastewater Reclamation Criteria (State of CA, 1978) for variety of water reuse applications. Based on the professional judgment and the research findings, California Department of Health Services was able to ensure safe and reliable water reuse practices. Lecture No: 26 26.1.Aerobic composting Composting is a biological process based on aerobic transformation of biodegradable wastes. The result of composting is a dark, humus like material that has fertilizing and soil texture improving properties. Composting can be used on almost all types of biodegradable wastes such as food residues, yard waste, and sewage sludge. During the composting process

oxygen is consumed, CO2, H2O and energy (heat) is produced. The overall reaction occurring during composting can in a simple manner be formulated as Biowaste + O2 → microbial biomass + non-degradable matter + CO2 + H2O + Heat Heat production causes the temperature in the composting material to rise and increases the biological degradation rate in the early stages of the composting process. Later when the easily degradable organic material has been degraded the rate of transformation and temperature gradually decreases to ambient levels. 26.2. Composting biology Most of the organisms responsible for the transformation of the organic material are microbes. Some important groups of microorganisms are bacteria, actinomycetes, and fungi. The bacteria are responsible for the turnover of approximately 80-90% of the organic matter transformed. Bacteria grow faster and are better adapted to the low oxygen concentrations and high temperatures often found in the early stages of the composting process. Actinomycetes are a group of filamentous organisms that are often found in blue-grey powder-like colonies. Both actinomycetes and fungi are relatively slow growing organisms that are less tolerant of low oxygen concentrations and high temperatures compared to the bacteria. The microbial populations and the temperature in the compost often follow a specific pattern dictated by the degradation of compounds in the organic matter. The composting process can be divided into four phases. The initial phase is the first period after initiation of the compost process where the temperature rises to about 50oC over a period of a few days (Fig. 26.1). During this phase the population of especially bacteria increases rapidly and compounds that are easily degradable, such as sugars, starch, proteins and fats are degraded. Due to the rapid rate of degradation and oxygen consumption it is often difficult to provide enough oxygen for the biological processes and the compost will have a tendency to develop anaerobic pockets. Modest decreases in pH may be observed due to the production of organic acids by anaerobic organisms. The organisms active during the initial phase are mesophilic (optimal temperature 35-45 oC) and thermophilic (optimal temperature (55 – 60 oC) bacteria.

Fig 26.1. Top: Microbial succession during the composting process Bottom: Idealized temperature variation in the compost during the course of the composting process If the conditions in the composting material are well maintained the composting material are well maintained the composting process will normally enter the thermophilic phase next. This phase involves especially thermophilic bacteria and also certain thermophilic actinomycetes and fungi. During this phase the temperature can exceed 70oC and temperatures as high as 80-85 oC have been observed during composting o sewage sludge. The pH usually increases to about 7.5 due to the destruction of the organic acids. Near the end of the thermophilic phase when the readily degradable organic material has been removed by the microorganisms only organic materials such as hemicellulose, lignin, chitin, and similar compounds that are more difficult to degrade remain. The microbial activity especially concerning the bacteria begins to decrease and the temperature in the compost begins to fall. At this point the composting process is not yet finished and the compost is sometimes called raw compost. Upon completion of the thermophilic phase the temperature decreases to levels where the mesophilic organisms have their optimum and the composting process enters the mesophilic phase. During this phase where the temperature ranges between 35 and 45 oC the more difficult to-degrade components such as cellulose and lignin are decomposed. During

the mesophilic phase several types of bacteria are still very active but it is especially the actinomycetes and fungi that are important during this phase. Actinomycetes and fungi are better adapted to utilize the more difficult degradable compounds compared to most of the bacteria. Some fungi can even produce penicillin that will kill some of the bacteria. The mesophilic phase can take up to several weeks to complete. At the end of the mesophilic phase the compost is often called finished compost. The final phase of the composting process is termed the cooling phase, during this phase the temperature slowly decreases to near ambient levels during a time span of several weeks, the microbial degradation of the organic material will be almost completed when entering the cooling phase and the rate of degradation will approach that of a natural soil. The organic matter remaining consists of very complex compounds with humus like structures that are difficult to degrade. The pH during this phase will normally stay relatively constant at about 8. Towards the end of the cooling phase higher organisms such as worms and insects will often colonize the compost. The compost is now termed mature compost and the structure of the organic matter in the compost will closely resemble that of humus.