Wastewater

C H A P T E R

13 Wastewater Daniel A. Vallero Pratt School of Engineering, Duke University, Durham, NC, United States

O U T L I N E 1. Introduction 1.1 Treatment Rationales

259 260

2. Water Pollution 2.1 Biological Agents 2.2 Inorganic Substances 2.3 Organic Substances

260 260 260 262

3. Wastewater Treatment 3.1 Solids Removal 3.2 Decreasing Oxygen Demand 3.3 Free Energy and Wastewater Treatment

263 264 265

4. Metabolic Mimicry

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274

6. Transport and Transformation

276

7. Environmental Fate

280

8. Wastewater Mechanisms 8.1 Monod Equation

283 287

9. Wastewater Data and Knowledgebases

288

References

289

Further Reading

290

268

1 INTRODUCTION Depending on the specific topic, waste management requires at least a modicum of understanding of wastewater. Arguably, the principal concern for waste managers is to minimize damage to air and water in whatever they do to handle solid waste or waste in any physical form. An understanding of wastewater must be part of the comprehensive and systematic waste management approach of this book. We would

Waste https://doi.org/10.1016/B978-0-12-815060-3.00013-X

5. Wastewater Kinetics

be remiss, within this systems perspective, not to address wastewater here (and air in Chapter 24). This is particularly important since many of techniques for addressing wastewater also apply to other wastestreams, especially environmental biotechnology for treating hazardous wastes, which will be addressed in more detail in Chapter 31. The environmental engineer understands that whatever the design, at some point in time and space, water needs to be protected. The

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Copyright # 2019 Elsevier Inc. All rights reserved.

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performance specifications during the active life and following closure of a waste site or landfill always include water protection. This may translate into limits on what and how much can leach into groundwater; what may be allowed to flow directly into a stream; or what should be allowed to flow indirectly onto soil, percolate as soil water, and eventually reach the water as a nonpoint source. These and countless other waste management scenarios include wastewater in one or more of the stages of design, operation, and decommissioning of waste facilities.

1.1 Treatment Rationales In the United States, water is treated for two basic reasons, water bodies and drinking water. Another way to view this distinction is that the Clean Water Act [1] is mainly concerned with addressing “dirty water” and the Safe Drinking Water Act is mainly concerned that “clean water” is delivered to the human population [2]. The Clean Water Act focuses on cleaning up impaired water bodies, for example, limits on the amounts and types of materials that can be discharged in to streams and other surface waters. The focus of the Safe Drinking Water Act is on the amount of specified chemical and biological substances that can be in a public water supply, that is, maximum contaminant levels (MCLs) [3]. Thus wastewater is the direct focus of the Clean Water Act and the indirect focus on the Safe Drinking Water Act.

2 WATER POLLUTION Wastewater can contain substances that are highly toxic to aquatic life and human populations. The pollutants may be dissolved or suspended in the water. Thus wastewater can either directly enter surface waters, such as through an outfall structure, or reach a storm sewer system. Once the polluted water reaches the surface water or groundwater, it harms both in many ways. The suspended matter will

increase turbidity, diminishing light infiltration and change the microbial populations in the receiving waters. As is always true in environmental systems, the effects are not limited to a single change. The reduction of light can also substantially diminish the growth capacity of aquatic plants and disrupts normal processes, such as seasonal turnover and nutrient cycling in the surface waters. Toxic contaminants, both organic and inorganic are carried by suspended particles and dissolved in the water, increasing the organic load and transporting hazardous chemical compounds. Laws and regulations prescribe which substances should be measured in public waters. In the United States, the EPA sets rules on which biological and chemical need to be measured and the methods needed to test them [4]. For the specific procedures for conducting these tests, the waste manager should consult the EPA website, “Clean Water Act Analytical Methods” [5].

2.1 Biological Agents A primary concern of water is whether it carries pathogenic microbes. When wastewater contains these organisms, they may reach drinking water supplies, recreation facilities, and other places where people may be exposed. They may also enter the food chain, where they can reside at various trophic levels and remain in reservoirs, for example, in fish, that are consumed. Table 13.1 presents some of the methods used to determine the extent of microbial contamination of waters.

2.2 Inorganic Substances Wastewater and sludge contain almost every element in the periodic chart. These exist in many forms. Chemists differentiate compounds as either organic or inorganic. If a compound lacks a carbon atom that is covalently bonded to another carbon or hydrogen atom, it is classified as “inorganic.” Inorganics are often further differentiated as metals and nonmetals.

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2 WATER POLLUTION

TABLE 13.1

Biological Methods for Wastewater and Sewage Sludge

Parameter

Method

1. Coliform (fecal), number per 100 mL or number per gram dry weight

Most probable number (MPN), 5-tube, 3-dilution, or Multiple tube/multiple well, or Membrane filter (MF), single step

2. Coliform (fecal) in presence of chlorine, number per 100 mL

MPN, 5-tube, 3-dilution, or MF, single step

3. Coliform (total), number per 100 mL

MPN, 5-tube, 3-dilution, or MF2, single step or two step

4. Coliform (total), in presence of chlorine, number per 100 mL

MPN, 5-tube, 3-dilution, or MF with enrichment

a

5. E. coli, number per 100 mL

MPN multiple tube, or Multiple tube/multiple well, or MF single step

6. Fecal streptococci, number per 100 mL

MPN, 5-tube, 3-dilution, or MF, or Plate count

7. Enterococci, number per 100 mL

MPN, 5-tube, 3-dilution, or MPN, multiple tube/multiple well, or MF single step or Plate count

8.Salmonella number per gram dry weight

MPN multiple tube

9. Toxicity, acute, fresh water organisms, LC50, percent effluent

Ceriodaphnia dubia acute Daphnia puplex and Daphnia magna acute Fathead Minnow, Pimephales promelas, and Bannerfin shiner, Cyprinella leedsi, acute Rainbow Trout, Oncorhynchus mykiss, and brook trout, Salvelinus fontinalis, acute

10. Toxicity, acute, estuarine and marine organisms of the Atlantic Ocean and Gulf of Mexico, LC50, percent effluent

Mysid, Mysidopsis bahia, acute Sheepshead Minnow, Cyprinodon variegatus, acute Silverside, Menidia beryllina, Menidia, and Menidia peninsulae, acute Continued

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262 TABLE 13.1

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Biological Methods for Wastewater and Sewage Sludge—cont’d

Parameter

Method

11. Toxicity, chronic, fresh water organisms, NOEC or IC25, percent effluent

Fathead minnow, P. promelas, larval survival and growth Fathead minnow, P. promelas, embryo-larval survival and teratogenicity Daphnia, C. dubia, survival and reproduction Green alga, Selenastrum capricornutum, growth

12. Toxicity, chronic, estuarine and marine organisms of the Atlantic Ocean and Gulf of Mexico, NOEC or IC25, percent effluent

Sheepshead minnow, C. variegatus, larval survival and growth Sheepshead minnow, C. variegatus, embryolarval survival and teratogenicity Inland silverside, Menidia beryllina, larval survival and growth Mysid, M. bahia, survival, growth, and fecundity Sea urchin, Arbacia punctulata, fertilization

a

Approved by the U.S. EPA for enumeration of target organism in wastewater effluent. Notes: When the membrane filter (MF) method has been used previously to test waters with high turbidity, large numbers of noncoliform bacteria, or samples that may contain organisms stressed by chlorine, a parallel test should be conducted with a multiple-tube technique to demonstrate applicability and comparability of results. U.S. Environmental Protection Agency, "Identification of test procedures," in 40 CFR 136.3, 2017.

Numerous methods are available for testing the inorganic features of wastewater [4]. These include acidity, ammonia, boron, bromide, calcium, chloride, cyanide, fluoride, nitrogen, nitrate-nitrite, phosphorus, potassium, silica, sodium, sulfate, sulfite, sulfides, and acid mine drainage. The methods for nonmetal detection vary widely, including titrimetric, ion chromatography, ion selective electrode, amperometric, spectrophotometric, digestion, distillation, and colorimetric. The principal metals, metalloids, and their compounds that often measured in wastewater and receiving waters include aluminum, antimony, arsenic, barium, beryllium, cadmium, chromium VI, and total, cobalt, copper, gold, iridium, iron, lead, magnesium, manganese, mercury, molybdenum, nickel, palladium, platinum, rhodium, ruthenium, selenium, silver, thallium, tin, titanium, vanadium, and zinc. Some of the methods for analyzing metals

include atomic absorption and inductively coupled plasma. Other measured parameters include acidity, color, hardness, hydrogen ion, specific conductance (indication of dissolved solids), temperature, turbidity, and even some parameters that are chemically organic, for example, surfactants, oil, and grease.

2.3 Organic Substances The U.S. organic compounds of concern in water as to whether they are pesticide ingredients [4]. The nonpesticide organics include those with widely ranging properties, including aqueous solubility and vapor pressure. This is because solubility is not the only and not necessarily the most important property for predicting the concentrations of organic compounds in wastewater and receiving waters. Recall that wastewater almost always has a high

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concentration of suspended solids and even highly hydrophobic compounds will sorb to these particles. The water may also be mixed with other solvents that allow for cosolvation; the process in which a substance is first dissolved in one solvent and then that solution is mixed with another solvent. This means that even a hydrophobic compound in wastewater can mix with the receiving water flow. For example, a dense nonaqueous phase liquid (DNAPL) like chlorinated benzene can migrate into and move within water bodies when it is first dissolved in an alcohol or an organic solvent that mixes with water (e.g., benzene or toluene). This may be particularly problematic for soils and groundwater, since the “D” in DNAPL means this substance is denser than water. Thus particularly if the organic solvent dissipates, the DNAPL will migrate downward in the soil water and groundwater because it has undergone cosolvation with the water. Likewise, the ordinarily lipophilic compound can be transported in the vadose zone or upper part of the zone of saturation where it undergoes cosolvation with water and a light nonaqueous phase liquid (LNAPL), for example, toluene. That is, the LNAPL can be the cosolvent with water, which allows other DNAPLs and LNAPLs to move away from the wastewater source. Since wastewater is a mixture, its constituents will partition within an environmental compartment at different rates. Processes like sorption and cosolvation increase the ease with which a dense substance that moves toward the bottom of the aquifer in with time may partition between lighter and denser constituents. The lighter constituents will be more likely to migrate upward and could become air pollutants via volatilization than the denser constituents of the mixture, which will likely remain below the surface. However, this should be considered along with other inherent properties and environmental conditions. Of course, some constituents may resist dissolution and continue to migrate downwardly. Therefore partitioning

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coefficients like Kow must be applied to each constituent and that partitioning is highly specific to the mixture and to environmental conditions [6, 7].

3 WASTEWATER TREATMENT The most common conception of the word “waste,” at least for environmental engineers, is solid waste. However, solid waste management often includes the identical or slightly modified versions of the wastewater technologies. Wastewater treatment involves physics, chemistry, and biology, that is, biochemodynamics. The type of wastewater treatment needed depends on the type of waste and the desired water quality after treatment. For example, the earliest stages of wastewater treatment are physical, that is, removing solids, such as by settling. The next stages also include physics, but also chemistry and biology, in order to understand the transforming of wastewater into higher quality water. Most municipal wastewater is treated by the screening of materials, sedimentation, and by a secondary treatment (see Fig. 13.1). Sedimentation removes the denser solids. Secondary treatment is usually biological, that is, environmental biotechnology, such as activated sludge or trickling filter. Often, these steps are followed by another, that is, tertiary treatment, which may include disinfection, nutrient removal, or additional solids removal beyond the sedimentation step [8]. The hazards in wastewaters can be direct, owing to its biological or chemical components, for example, pathogenic bacteria and heavy metals, respectively. The hazard may also be more indirect, such as the amount and types of solids and organic substances, which may induce problems like eutrophication of lakes or depletion of oxygen by enhancing the growth of otherwise nonpathogenic bacteria. Sometimes the contaminant can be both directly and

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FIG. 13.1

13. WASTEWATER

Prototypical wastewater treatment flowchart.

indirectly hazardous, for example, ammonia that kills fish directly or serves as a source of nitrogen that induces microbial growth, leading to eutrophication. Many of the same parameters important to solid waste managers are used in measuring and treating wastewater. These include solids, biochemical oxygen demand (BOD), fecal coliform, and xenobiotic chemicals.

3.1 Solids Removal Solid waste and wastewater are relative terms that depend on a gradient of water content. Wastewater has a much higher water content than sludge, which in turn has a higher water content than most landfill wastes. Conversely, wastewater has the lowest solids content. Solids harm surface waters by loading with organic

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matter and with toxic substances either sorbed to the suspended solids or as dissolved solids. Thus wastewater’s solids content must be measured to meet national, provincial, state, and/or local permit requirements. Solids also indicate the efficiency of process controls. For example, the percent of total solids (TS) indicates how much dewatering of sludge is needed before handling. An effluent permit may limit the percent of total suspended solids (TSS) that may be discharged to a stream. A wastewater treatment plant operator may measure the percent of total volatile solids (TVS) and total volatile suspended solids (TVSS) to indicate how well a digester or trickling filter is operating or to indicate proper food-to-mass (F:M) ratio loading rates for activated sludge treatment [9]. Wastewater solids are classified by size and mass. Total solids include, in descending size order, dissolved, suspended, or settleable solids. Solids content in wastewater is determined by evaporating the water and weighing the residue. In the case of TS, a known volume of the unfiltered water is heated at 103–105°C for 1 h [10]. The TS sample can then be filtered and evaporated to find the total dissolved solids (TDS) and TSS, that is, the solids that pass through the filter are TDS, and the solids that remain in and on the filter are TSS. These are further differentiated as fixed versus volatile solids. Fixed solids are those that remain after igniting the residue from TS. Volatile solids are those that are driven off by ignition. The actual oven and ignition temperatures, filter pore size and type, and other conditions are specified by governmental methods, for example, Method 1684 [10], or standards handbooks, especially Standard Methods for the Examination of Water and Wastewater [11]. Removal of solids is realized by physical processes, especially gravity, that is, denser particles settle. Chemical processes also aid in solids removal, such as adding chemical compounds that encourage smaller, lighter particles and colloids to aggregate and precipitate. Finally, biological processes are involved in solids

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removal. As mentioned, secondary treatment involves microbial degradation of organic matter. Engineering systems designed to achieve these processes are discussed in Chapter 31. These include activated sludge, aerated lagoons, and biofilters.

3.2 Decreasing Oxygen Demand The increased input of nitrogen, phosphorus, and other nutrients encourages the growth of algae, leading to algal blooms. When these algae die, they fall to the lake bottom. As they continue to decompose, more oxygen is demanded by the bacteria and abiotic processes, so the concentrations of dissolved oxygen (DO) decline. As a result, fish and other organisms at higher trophic levels in the food chain are stressed. This can result in fish kills and complete alterations to aquatic habitats. In addition, pathogenic microbes may be suspended or sorbed to suspended particles in the wastewater. Coliform like Escherichia coli and other microbes can reach water supplies, where they cause major human health problems. Waste managers must communicate with wastewater experts and decision makers, since they both have mutual objectives and may unwittingly be working at cross-purposes. When water pollutants are released from landfills, sludge drying beds, and other waste management processes, water systems received the brunt of impact. Organic compounds, nutrients, and other wastewater constituents “demand” oxygen. For example, leachate from waste facilities contains nutrients that, when they reach the water systems, affect the amount of dissolved oxygen (DO). The biota in these systems vary in their optimal ranges of growth and metabolism (e.g., algae add some O2 via photosynthesis, but use some O2 for metabolism, whereas the bacteria are net consumers of DO). Both solids content and BOD are commonly used by environmental engineers to quantify water pollution. The BOD is the amount of

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oxygen consumed with time. Most of the BOD is attributable to decomposing organic matter under aerobic conditions. However, since it is an empirical test, some of the demand may be “abiotic,” that is, chemical processes other than microbial. The BOD is measured by incubating a sealed sample of water with the wastewater or other water (e.g., water at points downstream from the outfall). The DO is measured before sealing the container, that is, at time ¼ 0 (t0). After sealing, the container is held for 5 days (t5), and the DO measured again. The loss of oxygen is determined by comparing the O2 concentration of the sample at (t0) to the concentration at (t5). Samples are commonly diluted before incubation to prevent the bacteria from depleting all of the oxygen in the sample before the test is complete [12]. Thus BOD5 is a simple calculation of measured DO at the beginning time, that is, the initial DO (D1), measured immediately after it is taken from the source minus the DO of the same water measured exactly 5 days after D1, that is, D5: BOD ¼

D1  D5 P

(13.1)

where P ¼ decimal volumetric fraction of water utilized. D units are in mg L1. If the dilution water is seeded, the calculation becomes: BOD ¼

ðD1  D5 Þ  ðB1  B5 Þf P

(13.2)

where B1 ¼ initial DO of seed control, B5 ¼ final DO of seed control, and f ¼ the ratio of seed in seed in D1 sample to seed in control ¼ % % seed in B1 . B units 1 are in mg L . For example, to find the BOD5 value for a 10 mL water sample added to 300 mL of dilution water with a measured DO of 7 mg L1 and a measured DO of 4 mg L1 5 days later: P¼ BOD5 ¼

10 ¼ 0:03 300

74 ¼ 100 mg L1 0:03

Thus the microbial population in this water is demanding 100 mg L1 dissolved oxygen over the 5-day period. So, if a conventional municipal wastewater treatment system is achieving 95% treatment efficiency, the effluent discharged from this plant would be 5 mg L1. Chemical oxygen demand (COD) does not differentiate between biologically available and inert organic matter, and it is a measure of the total quantity of oxygen required to oxidize all organic material completely to carbon dioxide and water. COD values always exceed BOD values for the same sample. COD (mg L1) is measured by oxidation using potassium dichromate (K2Cr2O7) in the presence of sulfuric acid (H2SO4) and silver. By convention, 1 g of carbohydrate or 1 g of protein accounts for about 1 g of COD. On average, the ratio BOD:COD is 0.5. If the ratio is <0.3, the water sample likely contains elevated concentrations of recalcitrant organic compounds, that is, compounds that resist biodegradation [13]. That is, there are numerous carbon-based compounds in the sample, but the microbial populations are not efficiently using them for carbon and energy sources. This is the advantage of having both BOD and COD measurements. Sometimes, however, COD measurements are conducted simply because they require only a few hours compared to the 5 days for BOD. Since available carbon is a limiting factor in the demand, the BOD reaches a threshold, known as the ultimate carbonaceous BOD. If carbonaceous compounds comprised the entire O2 demand, the curve would become a plateau (see Fig. 13.2). However, there are other sources of demand. Notably, microbial populations will continue to demand O2 from the water to degrade other compounds, especially nitrogenous compounds, which account for the bump in the BOD curve [14]. Thus in addition to serving as an indication of the amount of molecular oxygen needed for biological treatment of the organic matter, BOD also provides a guide to sizing a treatment process, assign its efficiency,

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BOD (mg L–1)

Nitrogenous oxygen demand

Ultimate carbonaceous BOD

Carbonaceous oxygen demand

BOD5 value

5

10

15

20

Time (days)

FIG. 13.2 Biochemical oxygen demand (BOD) curve showing ultimate carbonaceous BOD and nitrogenous BOD. Adapted from C. P. Gerba and I. L. Pepper, “Wastewater treatment and biosolids reuse," in: Environmental Microbiology, second ed. Elsevier Inc., New York, 2009, pp. 503–530.

and giving operators and regulators information about whether the facility is meeting its design criteria and is complying with pollution control permits. If effluent with high BOD concentrations reaches surface waters, it may diminish DO to levels lethal to some fish and many aquatic insects. As the water body reaerates as a result of mixing with the atmosphere and by algal photosynthesis, O2 is added to the water, the oxygen levels will slowly increase downstream. The drop and rise in DO concentrations downstream from a source of BOD is known as the DO sag curve, because the concentration of dissolved oxygen “sags” as the microbes deplete it. So, the falling O2 concentrations fall with both time and distance from the point where the high BOD substances enter the water. As discussed in Chapter 33, the amount of any substance, for example, nutrient, pollutant, or microbes, that is discharged into a system by point and nonpoint sources is known as the load.

Chapter 33 explains how nutrient loading, pollutant loading, and microorganism loading can be quantified as to the gross mass of the substance entering an ecosystem and by the response of the ecosystem to the load. Calculating these loads must account for background sources (sometimes referred to as “natural” sources). The linkages between biota and their environments are influenced by the availability and forms of nutrients. When rain containing nitrates falls on wetland water and land surfaces, for instance, plant roots take it up and metabolize it into organic forms of nitrogen. Meanwhile, bacteria in sediments reduce it to ammonia. In the opposite direction, the reduced forms are oxidized by other bacteria (e.g., Nitrosomonas converts ammonium ions [NH+4 ] to  nitrite [NO 2 ], Nitrobacter converts the NO2 to  nitrate [NO3 ]). Thus two opposite reactions are at play constantly in loading scenarios. In this instance, the ammonification (or deamination) and nitrification reactions are as follows:

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Ammonification : NH4+ + OH , NH3 + H2 O (13.3) + Nitrification : NH4+ + 2O2 ! NO 3 + 2H + H2 O

(13.4) Note that ammonification is written as an equilibrium reaction, and nitrification is an oxidation reaction. However, the loading depends on numerous environmental conditions, including pH. Thus under acidic conditions, most of the ammonia nitrogen will ionize to the ammonium ion and under basic conditions the nonionized ammonia concentrations will increase in proportion to the ammonium ion concentration. This is important since nonionized ammonia is very toxic to aquatic biota, whereas ionized species are nutrients needed for plants and algae growth. Add to this, the decomposition of organic matter, and the loading becomes a complicated mix of reactions: Organic nitrogen + O2 ! NH3 nitrogen + O2 ! NO2 nitrogen + O2 ! NO3 nitrogen + O2 (13.5) Even this is a gross oversimplification, since the oxidation and reduction depend on the types of bacteria present and other physicochemical properties of the water. For example, the O2 required for nitrification is theoretically 4.56mg O2 mg2 NH+4 , but this is an autotrophic reaction. Thus O2 is being produced by the nitrifying bacteria, decreasing the amount needed. However, the growth rate of nitrifying microbes is far less than that of heterotrophic microbes decomposing organic wastes. Therefore when large amounts of organic matter are being degraded, the nitrifiers’ growth rate will be sharply limited by the heterotrophs, which means the rate of nitrification will commensurately be decreased [15]. Furthermore, these same responses to conditions exist for all other nutrients in the system, for example, organic P will be oxidized and oxidized P species reduced, etc. Thus the kinetics of nutrient loading are complex.

3.3 Free Energy and Wastewater Treatment Both primary and secondary treatments decrease the organic load of the wastewater, that is, the BOD. However, it is the microbial degradation in secondary treatment that accounts for most of the BOD removal. This is accomplished by free energy, that is, the measure of a system’s ability to do work (this section is based on the free energy discussion in reference [14]). If reactants have greater free energy than the products in a chemical reaction, then energy is released from the reaction; that is, reaction is exergonic. Conversely, if the products from the reaction have more energy than the reactants, energy is consumed; that is, it is an endergonic reaction. Equilibrium constants are ascertained thermodynamically by employing the Gibbs free energy (G) change for the complete reaction expressed as: ΔG ¼ ΔH  TΔS

(13.6)

where ΔG is the energy liberated or absorbed in the equilibrium by the reaction at constant T. ΔH is the system’s enthalpy change and ΔS is its entropy change. Enthalpy is the thermodynamic property expressed as: H ¼ U + pV

(13.7)

where U is the system’s internal energy. The relationship between a change in free energy and equilibria can be expressed by: 0

ΔG∗ ¼ ΔG∗f + RT lnKeq

(13.8)

where ΔG∗0 f ¼ free energy of formation at steady state (kJ g mol1). Thus the total energy in systems is known as enthalpy (H) and the usable energy is known as free energy (G). Living cells need G for all chemical reactions, especially cell growth, cell division, and cell metabolism and health. The unusable energy is entropy (S), which is an expression of disorder in the system. Since disorder tends to increase because of the many conversion steps outside and inside of the cell, the

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cells have adapted ways of improving efficiencies. This is the key to finding biological treatment technologies. Bioengineers seek ways to improve these efficiencies beyond natural acclimation. Thus to understand both wastewater biotechnologies, the processes that underlie microbial metabolism must be characterized. All cells must carry out two very basic tasks to survive and grow. They must undergo biosynthesis, that is, they must synthesize new biomolecules to construct cellular components. They must also harvest energy. Metabolism consists

of the aggregate complement of the chemical reactions of these two processes. Thus metabolism is the cellular process that derives energy from a cell’s surroundings and uses this energy to operate and to construct even more cellular material. Energy that does chemical work is exemplified by cellular processes (see Fig. 13.3). Catabolism consists of reactions that react with molecules in the energy source, that is, incoming food, such as carbohydrates. These reactions generate energy by breaking down these larger

Catabolism

Anabolism Energy source

Cell Biomass Walls, membranes, and other cell structures

Energy

Energy

Proteins, nucleic acids— macromolecules

Amino acids, nucleotides—subunits

Energy Precursors Nutrients Waste products (CO2,organic acids)

(Compounds containing N, S,)

FIG. 13.3 Cellular metabolism results from catabolic reactions that break down compounds to gain energy that is used to build biomolecules (anabolic metabolism) from nutrients that are taken up by the cell, beginning with simple precursors, then subunits, macromolecules. From these biomolecules, the cellular structures are built. From: D. Vallero, Environmental Biotechnology: A Biosystems Approach. Elsevier Science, 2015.

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molecules. Anabolism consists of reactions that synthesize parts of the cell, so they require energy; that is, anabolic reactions use the energy gained from the catabolic reactions. Anabolism and catabolism are type of dynamic equilibrium. Anabolism synthesizes while catabolism destroys. However, anabolism can only build the cellular components by using the energy it receives from the catabolistic destruction of organic compounds. Thus as the cell grows, its food supply (organic matter in the wastewater) declines. The efficiency of wastewater treatment depends on the microbial population’s ability to use these two metabolic functions to breakdown organic substrates. Thus as shown in Fig. 13.4, the type of wastewater determines the best bacterial species and the conditions needed. Ideally, the microbes in the media in a trickling filter, for example, use the organic compounds as their exclusive source of energy (catabolism) and their sole source of carbon (anabolism). Microbes, for example, algae, bacteria, and fungi, are essentially miniature and efficient chemical factories that mediate reactions at various rates (kinetics) until equilibrium is reached.

These “simple” organisms (and complex organisms alike) need to transfer energy from one site to another to power their machinery needed to stay alive and reproduce. Microbes play a large role in degrading pollutants, whether in natural attenuation, where the available microbial populations adapt to the hazardous wastes as an energy source, or in engineered systems that do the same in a more highly concentrated substrate. Some of the biotechnological manipulation of microbes is aimed at enhancing their energy use or targeting the catabolic reactions toward specific groups of food, that is, organic compounds. Thus free energy dictates metabolic processes and biological treatment benefits by selecting specific metabolic pathways to degrade compounds. This occurs in a step-wise progression after the cell contacts the compound. The initial compound, that is, the parent, is converted into intermediate molecules by the chemical reactions and energy exchanges shown in Fig. 13.4 These intermediate compounds, as well as the ultimate end products can serve as precursor metabolites.

FIG. 13.4 Microbial oxidation that occurs during the degradation of organic compounds. Both the catabolic and anabolic processes generate oxidation products.

CO2 + water

m

ida

tio

bolis

Ox

Cata

n

Oxidation Oxidation products Products

Organic matter Organic Matter (substrate) (Substrate)

Ce

ll s

yn

Anab

olism

the

sis

New cells New Cells

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Endogenous respiration

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4 METABOLIC MIMICRY

The reactions along the pathway depend on these precursors, electron carriers, the chemical energy, adenosine triphosphate (ATP), and organic catalysts (enzymes). The reactant and product concentrations and environmental conditions, especially pH of the substrate, affect the observed Δ G* values. If a reaction’s Δ G* is a negative value, the free energy is released and the reaction will occur spontaneously, and the reaction is exergonic. If a reaction’s Δ G* is positive, the reaction will not occur spontaneously. However, the reverse reaction will take place, and the reaction is endergonic. Time and energy are limiting factors that determine whether a microbe can efficiently mediate a chemical reaction, so catalytic processes are usually needed. Since an enzyme is a biological catalyst, these compounds (proteins) speed up the chemical reactions of degradation without themselves being used up. They do so by helping to break chemical bonds in the reactant molecules (see Fig. 13.5). By decreasing the activation energy needed, a biochemical reaction can be initiated sooner and easier than if the enzymes were not present. Indeed, enzymes play a very large part in microbial metabolism. They facilitate each step along the metabolic pathway. As catalysts, enzymes reduce the reaction’s activation energy, which is the minimum free energy required for a molecule to undergo a specific reaction. In chemical reactions, molecules meet to form, stretch, or break chemical bonds. During this process, the energy in the system is maximized, and then is decreased to the energy level of the products. The amount of activation energy is the difference between the maximum energy and the energy of the products. This difference represents the energy barrier that must be overcome for a chemical reaction to take place. Catalysts (in this case, microbial enzymes) speed up and increase the likelihood of a reaction by reducing the amount of energy, that is, the activation energy, needed for the reaction. Enzymes are usually quite specific. An enzyme is limited in the kinds of substrate that it will

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catalyze. Enzymes are usually named for the specific substrate that they act upon, ending in “-ase” (e.g., RNA polymerase is specific to the formation of RNA, but DNA will be blocked). Thus the enzyme is a protein catalyst that has an active site at which the catalysis occurs. The enzyme can bind a limited number of substrate molecules. The binding site is specific, that is, other compounds do not fit the specific three-dimensional shape and structure of the active site (analogous to a specific key fitting a specific lock). The complex that results, that is, the enzyme–substrate complex, yields a product and a free enzyme. The most common microbial coupling of exergonic and endergonic reactions (see Fig. 13.6) by means of high-energy molecules to yield a net negative free energy is that of the nucleotide, ATP with Δ G* ¼  50 to 60kJ mol1 (12 to 15 kcal mol1). Other high-energy compounds also provide energy for reactions, including guanosine triphosphate, uridine triphosphate, cytosine triphosphate, and phosphoenolpyruvic acid. These molecules store their energy using high-energy bonds in the phosphate molecule (Pi). An example of free energy in microbial degradation is the possible first step in acetate metabolism by bacteria: Acetate + ATP ! acetyl  coenzyme A + ADP + Pi (13.9) In this case, the Pi represents a release of energy available to the cell. Conversely, to add the phosphate to the two-Pi structure ADP to form the three-Pi ATP requires energy (i.e., it is an endothermic process). Thus the microbe stores energy for later use when it adds the Pi to the ATP.

4 METABOLIC MIMICRY Many wastewater treatment technology classes have been renamed “environmental biotechnology,” since they mimic and enhance natural biological systems to treat wastewater.

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Ac tiv e

sit e

Matching substrates

Enzyme

Nonmatching substrate

Enzyme-substrate complex

Product

Enzyme

FIG. 13.5 Substrates specific to the enzyme. The breakdown of the enzyme–substrate complex yields a product and the enzyme becomes available for another catalytic reaction. The nonmatching substrate cannot enter into a complex, so it is not affected by the presence of this particular enzyme. However, another enzyme may have the “lock” to match this substrate’s “key.”

Historically, the same species found in soil, manure, and other natural substrates have been acclimated to use certain organic compounds in a waste product as energy sources by three specific mechanisms separately or simultaneously [16, 17]:

• use of the compound as an electron acceptor; • use of the compound as an electron donor; or • cometabolism. Use of a compound as an electron acceptor takes place under reducing conditions. Such

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4 METABOLIC MIMICRY

Energy

Activation energy (without catalyst)

Activation energy (with catalyst)

Reactants

Heat released to environment

Products

Direction of exothermic reaction

Energy Activation energy (without catalyst)

Activation energy (with catalyst) Absorbed heat

Products

Reactants

Direction of endothermic reaction

FIG. 13.6

Effect of a catalyst on an exothermic reaction (top) and on an endothermic reaction (bottom).

biotransformation needs a source of carbon (i.e., the electron donor) for microbial growth and metabolism. For example, if the compound contains chlorine, then reductive dehalogenation will occur. That is, the halogens that are bound to carbon and other elements are removed, usually reducing the toxicity of the compound. The electron donor carbon can be gained from various sources, both from natural and anthropogenically derived organic matter.

During electron donation, the compound is used as the primary substrate, so that the microbe obtains its energy and organic carbon from the compound. This may occur either under aerobic and under some anaerobic conditions. Lesser oxidized compounds (e.g., vinyl chloride, dichlorethylene, or 1,2-dichloroethane) are amenable to this mode of biotransformation. What is actually occurring is known as “cometabolism,” that is, the interaction

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“between enzyme specificity and metabolic regulation, the metabolic interdependence of microorganisms, and co-substrate requirements in the catabolism of xenobiotic compounds” [18]. Thus the degradation is catalyzed by an enzyme that is fortuitously produced by the organisms for other reasons of survival. The microbe receives no direct benefit from the degradation of the compound and may indeed be harmed or its growth and metabolism be inhibited during cometabolism, but if other energy sources are absent, the microbe is coaxed into using the toxic substance [14]. Of course genetic engineering and synthetic biology have pushed beyond acclimation to generating new strains of microbes tailored to treat specific wastes. Like wastewater treatment, engineers must find ways to treat solid waste, both abiotically and biotically. For example, a biofilter can treat a specific organic pollutant in solid wastes and wastewater, by combining physical methods, for example, pumping and air sparging, with biotechnological methods. Some processes are exclusively abiotic, for example, thermal treatment, but their residue may call for biotechnological treatment, for example, land application in which the elements in the residue are used as source of nutrients for plants and soil bacteria.

quantity that occurs with time, the wastewater reaction rate at which constituents of wastewater contaminants change into new chemical compounds can be written as: Reaction rate ¼

Change in product concentration Corresponding change in time (13.10)

and Reaction rate ¼

Change in reactant concentration Corresponding change in time (13.11)

When a compound breaks down in nature or in a wastewater treatment system, the change in product concentration drops proportionately with the reactant concentration. Thus for reactant substance A the kinetics is: Rate ¼ 

A kinetic reaction is denoted by the onedirectional arrow (!). In the environment, many processes are incomplete, such as the common problem of incomplete combustion in an incinerator and the generation of new compounds other than the two that result from complete oxidation, that is, carbon dioxide and water. The kinetics of a reaction refers to the rate at which reactants are transformed into products. Wastewater kinetics can be described as abiotic or by biological systems, such as microbial metabolism. Since a rate is a change in a

(13.12)

The negative sign denotes that the reactant concentration, that is, the parent contaminant in the wastewater, is decreasing. Thus the degradation product C resulting from the concentration increases in proportion to the decreasing concentration of the contaminant A: Rate ¼

5 WASTEWATER KINETICS

ΔðAÞ Δt

ΔðCÞ Δt

(13.3)

By convention, the concentration of the chemical is shown in parentheses to indicate that the system is not at equilibrium. The difference between an initial concentration and a final concentration, Δ(X): ΔðXÞ ¼ ΔðXÞfinal  ΔðXÞinitial

(13.14)

Thus the rate of reaction at any time is the negative of the slope of the tangent to the concentration curve at that specific time (see Fig. 13.7). For a reaction to occur, the molecules of the reactants must collide. The intensity of these collisions is expressed as the “rate law.” For example, when reacting A and B to yield product C

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5 WASTEWATER KINETICS

275

FIG. 13.7 The rate of reaction at any time is the negative of the slope of the tangent to the concentration curve at that time. The rate is higher at t1 than at t3. This kinetic rate is concentration dependent (first order). From: D. A. Vallero, Fundamentals of Air Pollution, fifth ed. Waltham, MA: Elsevier Academic Press, 2014, p. 999; J. N. Spencer, G. M. Bodner, and L. H. Rickard, Chemistry: Structure and Dynamics. New York: John Wiley & Sons, 2010.

(i.e., A + B ! C), the reaction rate increases in accordance with the increasing concentration of either A or B. If the amount of A is tripled, then the rate of this whole reaction triples. Thus the rate law for such a reaction is: Rate ¼ k½A½B

(13.15)

For the reaction X + Y ! Z, where the rate is only increased if the concentration of X is increased (changing the Y concentration has no effect on the rate law), the rate law must be: Rate ¼ k½X

(13.16)

According to Eqs. 13.10 and 13.11, the concentrations in the rate law are the concentrations of reacting chemical species at any specific point in time during the reaction. The rate expresses how fast reaction is happening at that time. The constant k in the equations is the rate constant, which is unique for every chemical reaction and is a fundamental physical constant for a reaction, depending on the environmental conditions (e.g., pH, temperature, pressure, type of solvent). The rate constant is the rate of the reaction when all reactants are present in a 1 M (M) concentration. Accordingly, the rate constant

k is the rate of reaction under conditions standardized by a unit concentration. By drawing a concentration curve for a contaminant that consists of an infinite number of points at each instant of time, an instantaneous rate can be calculated along the concentration curve. At each point on the curve the rate of reaction is directly proportional to the concentration of the compound at that moment in time. This is a physical demonstration of kinetic order. The overall kinetic order is the sum of the exponents (powers) of all the concentrations in the rate law. So for the rate k[A][B], the overall kinetic order is two. Such a rate describes a second-order reaction because the rate depends on the concentration of the reactant raised to the second power. Other decomposition rates, such as k[X], are first-order reactions because the rate depends on the concentration of the reactant raised to the first power. The kinetic order of each reactant is the power that its concentration is raised in the rate law. So, k[A][B] is first order for each reactant and k[X] is first-order X and zero-order for Y. In a zeroorder reaction, compounds degrade at a constant rate and are independent of reactant concentration.

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6 TRANSPORT AND TRANSFORMATION The waste manager’s immediate needs are usually local but may also extend beyond the agency or company. Scale is a key consideration when building a water knowledgebase [19]. To varying extents, water is part of every important physical, chemical, and biological mechanism and process at every scale from planetary to molecular. Water is a transport medium from every scale from intracellular, to the smallest water body, to an ecosystem, to a continent, and to the globe [20]. Fig. 13.1 depicts the “reactors” in the environment [21]. Streams, lakes, aquifers, and other water systems provide the means by which substances enter, change, and exit these reactors. The paths include a range of scales, such as antimicrobial chemicals released into the environment before being taken up by animal and human populations [14]. This involves the cellular scale, in which the chemical then reaches the organism (animal “reactor”), wherein the gut microbes build resistance to even new antimicrobials. The water bodies act at a higher scale (watershed “reactor”), spreading the resistant bacteria and decreasing the effectiveness of antibiotics needed to treat human diseases [21]. In fact, studies have shown that even before a human antibiotic reaches the marketplace, it may have lost substantial effectiveness given that chemically similar veterinary antibiotics may have already passed through Level II in Fig. 13.8, that is, in animals and the pathogenic bacteria have become resistant to the new chemical [21–24]. The waste manager must determine which pollutants will be affected, positively or negatively, by a proposed action, for example, designing and siting a landfill or an incinerator. Choosing the best alternative requires ways to estimate and project what will happen to these pollutants by the available actions and comparing these options by modeling how each contaminant’s behavior in the environment. Such

modeling is highly complex and complicated since biological and chemical contaminants can be transported by myriad processes at all scales before reaching human and ecosystem receptors (see Fig. 13.9) [13, 14]. A “parent compound” is transformed into degradation products, sometimes called “chemical daughters” or “progeny.” For example, pesticide kinetics often concerns itself with the change of the active ingredient in a pesticide to its degradation products. The degradation products themselves often undergo transformations, depending on environmental conditions. For example, as shown in Fig. 13.9, a parent compound that is highly reactive may degrade within a few minutes after exiting a stack, producing a less reactive degradation product (DP1). Downstream from the effluent discharge, water and compounds within it react with the degradation product, degrading 99% of it within a few hours. These newly transformed compound (DP2) may be less reactive and do not undergo further transformation. At distance A in Fig. 13.10, the measurements would indicate about equal mass of the parent and DP1. However, at distance B, none of the parent compound is detected, and about the same amount of DP1 is measured at distance A because there has been sufficient time since the peak concentration, for this transformation to take place. Very little of the DP2 is present at distance B since it only began to be produced since DP1’s degradation. At distance C, sufficient time has lapsed to allow all of mass of DP1 to be degraded into the less reactive DP2. Note that DP2 is not showing signs of decreasing even though all of DP1 is degraded, indicating it is not nearly as reactive as DP1. Indeed, it may travel a long distance before deposition. Persistent compounds can travel many miles before being degraded, volatilized, or deposited [20]. Incidentally, a compound, dissolved in water vapor or sorbed to solids, can move advectively with the stream. If this were the case for the three compounds in Fig. 13.10, the stream

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6 TRANSPORT AND TRANSFORMATION

277

Antimicrobial use Reactor Level 1

Animal microbial populations

Human microbial populations

Microbes introduced

Level 2

Confined feeding operations, aquaculture, farms, etc.

Microbial

Level 3

Wastes, effuents, emissions, drift

Microbial

Level 4

Ground & surface waters

into the environment

Healthcare facilities, longterm care, daycare centers, etc.

genetic mixing

Wastewater treatment plants, sewers, septic tanks, etc.

genetic mixing

Soil & sediments

FIG. 13.8 Reactors in the environment. In this case, the change in microbial resistance as a result of genetic mixing at four levels of animal and human reservoirs. Adapted from: F. Baquero, J.-L. Martı´nez, and R. Canto´n, "Antibiotics and antibiotic resistance in water environments," Curr. Opin. Biotechnol. 19(3), (2008), pp. 260–265.

concentrations would decrease to nearly zero if all of the solids settle into the sediment. Also, the transformations occur simultaneously in the water, sediment, and organisms. For example, a fraction of DP1 may be broken down abiotically (e.g., hydrolysis) or biotically (microbial metabolism) at distance A, but perhaps between A and B, most of the DP1will have been absorbed by fish and then stored in their tissue. A portion of DP1 may then convert to DP2 or

even a different degradation product (DP3). The DP1 is no longer available to the water, but continues to degrade and be stored in the fish, after which it may be a food source for other organisms and move through the food chain. The previous scenario highlights one of the values of taking deposition measurements, that is, to see what was in the atmosphere at the time of the rain event. Another value to see what is reaching the earth’s surface, since the air

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13. WASTEWATER

Stratosphere –

Ozone layer depletion

Troposphere – CFCs, CO2, CH4

Increased UV solar radiation Global warming

Relatively long atmospheric lifetimes: CFCs, CO2, CH4, persistent organic pollutants (including dioxins and PCBs) Hg, fine aerosols (PM2.5)

Reactions with –OH

Relatively short atmospheric lifetimes: SOx, NOx, CO, volatile organics, PM10 aerosols, heavy metals, high molecular weight organic compounds (not sorbed to fine aerosols), hydro-CFCs

VOCsx,NOx O3

SOx, NOx.... H2O Anthropogenic sources

Gas Natural sources

Acid precipitation

Deposition to terrestrial surfaces

Particulate matter Dry deposition

Runoff and snow melt

Producers

Consumers

Direct deposition to water, snow

Surface

Terrestrial food webs

Aquatic food webs Producers

Solutions Suspensions

Humans Particle sedimentation Decomposers

Wet (rain, snow) deposition

Consumers

Humans Decomposers

Sediment burial

FIG. 13.9 Pollutant transport and transformation processes. From (2003). The Sound Management of Chemicals (SMOC) Initiative of the Commission for Environmental Cooperation of North America: Overview and Update. Available from: http://www3.cec.org/ islandora/en/item/2031-overview-and-update-sound-management-chemicals-smoc-initiative-en.pdf. Adapted in D. Vallero, Environmental Contaminants: Assessment and Control. Amsterdam/Boston: Academic Press, 2010.

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6 TRANSPORT AND TRANSFORMATION

FIG. 13.10 Hypothetical transformation of a reactive compound (parent), followed by the transformation of its first degradation product (DP1) in a plume moving downstream from the point of effluent discharge. DP1 is further degraded into a less reactive chemical species (DP2), indicated by its steady state after all of DP2 has been transformed. DP2 will remain in the water until it is broken down, volatilized, and deposited. Adapted from: D. A. Vallero, Fundamentals of Air Pollution, fifth ed. Waltham, MA: Elsevier Academic Press, 2014, p. 999.

pollutants can become water or soil pollutants and can harm biota, for example, sulfate compounds that stress vegetation. The combination of deposition and transformation is an example of how physics and chemistry must be considered in concert to determine the severity of air pollution. In addition, equilibrium reactions also occur in the environment. Of course, getting to equilibrium requires a phase in which the reaction is dominated by kinetics. So, upon reaching equilibrium, the kinetic reactions (one-way arrows) would be replaced by two-way arrows ($). Changes in the environment or in the quantities of reactants and products can invoke a change back to kinetics. Incomplete reactions are very important sources of environmental contaminants. For example, these reactions generate products of incomplete combustion (PICs), such as CO, PAHs, dioxins, furans, and hexachlorobenzene (HCB). However, even the products of complete combustion are not completely environmentally acceptable. Both carbon dioxide and water are greenhouse gases. They are both essential to life on earth, but excessive amounts of CO2 are strongly suspected of altering climate, especially increasing the mean surface temperatures on earth. Thus CO2 is considered by many to be an “air pollutant.” In reality, water pollution results from complex interactions of substances in various states in the many environmental media where

substances are transformed. These transformations may render a compound that was previously not very toxic into a very toxic substance. Conversely, what may have been a precursor to an air pollutant may have been changed into a less toxic compound or compounds. During transport, a pollutant may also undergo chemical changes. These changes may form toxic compounds or other types of problems, for example, they may become stronger greenhouse gases such as CH4 formed by reduction of CO2. After deposition, chemical reactions occur in the soil, water, and biota. Thus pollutants are a result of a mix of chemical reactions each occurring at its own rate. Pollution seldom occurs in a single system. Indeed, a better way to consider these mixed systems is pseudo orders in the rate laws, that is, “pseudoopen” and “pseudo-closed” systems. Applying the qualifier “pseudo” to absolute terms is a good way to describe the actual rates that occur in the environment, as opposed to wellcontrolled laboratory experiments. The rate law for a chemical reaction is an equation that links the rate of the reaction with concentrations of reactants and rate constants (commonly, partial order reactions and rate coefficients). The rate is expressed as: r ¼ k½Ax  ½By

(13.17)

where k is the rate coefficient, the concentration of each chemical species is in brackets and the

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exponents are derived experimentally. A firstorder reaction occurs at a linear rate: r¼

d½A ¼ k½A1 ¼ k½A dt

(13.18)

Similarly, a second-order reaction rate is: r¼

d½ B ¼ k½B2 dt

(13.19)

or r¼

d½B ½ B 2

¼ kdt

(13.20)

The rate equations for third and next orders would follow the same format. In air pollution, numerous reactants are involved in reactions. Thus mixed second-order rate reactions can be expressed as: r¼

d½ B ¼ k½B½A dt

(13.21)

When reaction rates exceed first order, the chemical concentrations can be adjusted so that the kinetics appears to be first order. So, for the simple reaction 2A + B ! C, the rate law would be: 

1 d½A ¼ k½A½B 2 dt

(13.22)

Since the right side is a product of two reactants, the overall rate law is second order, but with excess concentrations of B, then [B] is almost completely unchanged in the reaction, that is [B] changes very much more slowly than does [A]. In this instance, Eq. 13.22 can be better written as: 

1 d½A  k0 ½A½B 2 dt

(13.23)

where k0 is a pseudo-first-order rate coefficient (units are inverse time, s1).

7 ENVIRONMENTAL FATE After the processes and mechanisms of transport and transformation, the pollutant reaches its fate; this can be expressed in terms of physical

and chemical properties together with the location of the substance, that is, what it becomes and the site where it remains. The “it” may be the released pollutant or it may be a degradation product, that is, a degradate. For example, the municipal wastewater and landfills receive pharmaceuticals and personal care products (PCPs) in their wastestreams [25]. The degradates of these chemicals can be more concentrated than the parent substances that the residents use. For example, the antiinflammatory drug, ibuprofen, and its metabolites, for example, carboxyibuprofen and hydroxyibuprofen have been measured in influent and effluent samples in municipal wastewater. Carboxyibuprofen has been found at higher concentrations than ibuprofen and hydroxyibuprofen in the influent [26]. Some of the transformation takes place within the human who ingests the drug, so what enters the wastewater is about 15% ibuprofen, 43% carboxyibuprofen, and 26% hydroxyibuprofen. The degradation continues in the environment. Hydroxyibuprofen has been found at higher concentrations than the other two chemicals in seawater. Similar results have been shown for other over-thecounter pharmaceuticals and PCPs [27]. Within a geologic timeframe, the environmental fate is not permanent. Indeed, a chemical may have many fates, since a substance undergoes numerous changes in location and form before reaching it final fate. With changes in conditions, this fate will also change. This new fate also becomes a release. A wastewater pollutant’s fate may be the sediment at a stream bottom, but with changes in hydrological conditions like a hurricane or a dredging project, the pollutant may be rereleased from the sediment, transported in the river, and find its way to an alluvial aquifer, that is, its new fate is the groundwater. The fate may again change if the aquifer is a source of drinking water, for example, a public water supply, where the water is drawn and treated. The treatment may be complete, so that the pollutant’s new fate is

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7 ENVIRONMENTAL FATE

the treatment sludge. If incomplete, the amount of pollutant that remains untreated may be transported in the water supply to consumers. Within the body of the consumer, the pollutant may undergo toxicokinetics, that is, absorption, distribution, metabolism, and excretion. The amount that is absorbed, but not excreted stays within the body, that is, the body burden. Thus the new fate is a person’s tissues. Of course, body burden has a time function. A very persistent compound may remain in tissue for years, whereas a more reactive compound may change very rapidly, for example, a few hours. For example, if a person is exposed to a chemical at 225 ng 100 kg1 year1 from drinking, 25 ng 100 kg1 year1 from ingesting food, and 50 ng 100 kg1 year1 from contact through skin, the person’s total accumulation of for the 20-year period would be (225 + 25 + 50)ng  20 ¼ 7000 ng. This person’s body would have built up 7 μg of the chemical over this 20-year span. As mentioned, the chemical dose during that period must not only include the absorption rate, but the distribution, metabolism, and elimination rates. If 80% of ingested chemical is absorbed by the gastrointestinal tract [28], with 90% of this amount cleared from the person’s body each year, the net amount is the product of the accumulated amount times the absorption rate times the remaining amount after clearance (i.e., 1—clearance rate): 7μg  0:8  ð1  0:9Þ ¼ 0:56μg Thus the person has 0.56 μg per 100 kg, that is, an accumulation concentration of 0.0056 μg or 5.6 mg of the chemical body burden after 20 years of exposure. This calculated body burden would increase if the absorption rate increased. The body burden would decrease if the clearance rate increased. For hydrophobic and lipophilic compounds, those with large octanol–water partitioning coefficients (Kow) like DDT (log Kow ¼ 6.9) the fate is likely to be in lipid-rich tissues. Thus the clearance rate could be decreased in proportion to the amount

281

of a compound that is stored in lipids, that is, lipophilic compounds generally have much longer half-lives than hydrophilic compounds in mammals. Conversely, the clearance rate of a completely reactive, nonbioaccumulating compound would be 100%, that is, none of the compound is stored. However, this does not necessarily mean that the person has no risk. Firstly, many nonpersistent compounds are highly and acutely toxic, so the long-term exposure to a toxic compound could be causing incremental harm. Secondly, even if the body burden of the parent compound is zero, its metabolites may accumulate. Metabolism must also be considered in fate calculations. The person exposed to DDT would not only have a body burden of the parent DDT, but also of DDT’s metabolites, dichlorodiphenyldichloroethylene (DDE) and dichlorodiphenyldichloroethane (DDD) which, like DDT, also persist and bioaccumulate in the body. It is not unusual for the metabolite body burden to exceed that of the parent compound. Within or outside of the body, a parent compound can undergo metabolism by various processes such as reduction, hydrolysis, or dehalogenation and depending on each metabolite having its own toxicity (see Fig. 13.11). Once formed, the new compounds will differ in severity and type of toxicity, for example, DDE’s mechanisms and/or rate in disrupting the endocrine system differ from DDT (i.e., antiandrogenic). Indeed, the metabolites of a nontoxic compound may become toxic by a process known as bioactivation. For example, most polycyclic aromatic hydrocarbons (PAHs), molecularly flat compounds with repeating benzene structures, are highly hydrophobic and difficult for an organism to eliminate (since most blood and cellular fluids are mainly water). One of the most toxic PAHs is benzo(a)pyrene, which is found in cigarette smoke, combustion of coal, coke oven emissions, and numerous other processes that use combustion. In the body, benzo (a)pyrene can react to produce an epoxide

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FIG. 13.11

Two degradation pathways for dichlorodiphenyltrichloroethane (DDT): top. Right: Reductive dechlorination to form dichlorodiphenyldichloroethane (DDD). Left: Elimination of HCl (H2 addition) to form dichlorodiphenyldichloroethylene (DDE). From: D. A. Vallero, Fundamentals of Air Pollution, fifth ed. Waltham, MA: Elsevier Academic Press, 2014, p. 999.

compound that is even more toxic and carcinogenic than the original benzo(a)pyrene (see Fig. 13.12). The fate can be gradual or may change suddenly in the case of natural disasters and human failure. Wastewater releases, under disaster

situations, can overcome an otherwise welldesigned system. Thus the waste manager and engineer must ensure that the uncertainties and complexities posed by a natural disaster can be contained by designing, siting, constructing, and operating their systems to be sufficiently

FIG. 13.12

Bioactivation of benzo(a)pyrene to form the carcinogenic metabolite benzo(a)pyrene-7,8-dihydrodiol9,10-epoxide. During metabolism, the biological catalysts (enzymes) cytochrome P-450 and epoxide hydrolase render greater polarity to the molecule, and in the process form diols and epoxides. These metabolites are more toxic than the parent compound.

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8 WASTEWATER MECHANISMS

resilient. To prepare for human failure, the engineered systems must provide for a sufficient and optimal operational timeframe, in which the failure rate is minimized, that is, the waste management system is reliable [29]. Reliability is discussed in greater detail in Chapter 36.

8 WASTEWATER MECHANISMS As mentioned, treating wastewater requires a knowledge of physical, chemical, and biological processes. Indeed, wastewater treatment often emulates natural processes, such as decay by microbes in water and soil. Certainly, dissolution of a chemical in water is a dominant process, but as mentioned, many of the compounds of concern are hydrophobic under laboratory conditions, but are often found in wastewater. Thus cosolvation and sorption become the crucial mechanisms at work in wastewater and sludge. Indeed, sorption is arguably the most important transfer process that determines the concentrations of many compounds in surface waters. The physicochemical transfer [30] of chemical is an interaction of the solute (i.e., the chemical being sorbed) with the surface of a solid surface and can be complex depending on the properties of the chemical and the water. Other fluids are often of such small concentrations that they do not determine the ultimate solid–liquid partitioning. While it is often acceptable to consider “net” sorption, let us consider briefly the four basic types or mechanisms of sorption: (a) Adsorption is the process wherein the chemical in solution attaches to a solid surface, which is a common sorption process in clay and organic constituents in soils. This simple adsorption mechanism can occur on clay particles where little carbon is available, such as in groundwater. (b) Absorption is the process that often occurs in porous materials so that the solute can diffuse into the particle and be sorbed onto

283

the inside surfaces of the particle. This commonly results from short-range electrostatic interactions between the surface and the contaminant. (c) Chemisorption is the process of integrating a chemical into a porous materials surface via a chemical reaction. In soil, this is usually the result of a covalent reaction between a mineral surface and the contaminant. (d) Ion exchange is the process by which positively charged ions (cations) are attracted to negatively charged particle surfaces or negatively charged ions (anions) are attracted to positively charged particle surfaces, causing ions on the particle surfaces to be displaced. Particles undergoing ion exchange can include soils, sediment, airborne particulate matter, or even biota, such as pollen particles. Cation exchange has been characterized as being the second most important chemical process on earth, after photosynthesis. This is because the cation exchange capacity (CEC), and to a lesser degree anion exchange capacity (AEC) in tropical soils, is the means by which nutrients are made available to plant roots. Without this process, the atmospheric nutrients and the minerals in the soil would not come together to provide for the abundant plant life on planet earth [31]. Both physics and chemistry are at work in all four types, all of which are important in wastewater treatment, since they function at surfaces and are crucial to biofilm and molecular exchanges; such exchanges occur in trickling filter systems. The first two types of sorption are predominantly controlled by physical factors, and the second two are combinations of chemical reactions and physical processes. Generally, sorption reactions in biochemodynamic systems involve the three processes [32]: 1. The chemical contaminant’s transport in water due to distributions between the aqueous phase and particles.

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2. The aggregation and transport of the contaminant as a result of electrostatic properties of suspended solids. 3. Surface reactions such as dissociation, surface catalysis, and precipitation of the chemical contaminant. When a contaminant enters soil, some of the chemical remains in soil solution and the remainder is adsorbed onto the surfaces of the soil particles. Sometimes this sorption is strong due to cations adsorbing to the negatively charged soil particles. In other cases the attraction is weak. Sorption of chemicals on solid surfaces needs to be understood because they hold onto contaminants, not allowing them to move freely with the pore water or in the soil solution. Therefore sorption slows the rate at which substances move downwardly through the soil profile.

Biomolecules and xenobiotic compounds eventually establish a balance between the mass on the solid surfaces and the mass that is in solution. Molecules will migrate from one phase to another to maintain this balance. The properties of both the chemical and the soil (or other matrix) will determine how and at what rates the molecules partition into the solid and liquid phases. These physicochemical relationships, known as sorption isotherms, are found experimentally. Fig. 13.13 illustrates three isotherms for an organic compound, using three different substrates [35]. The x-axis in Fig. 13.13 shows the concentration of pyrene dissolved in water, and the y-axis shows the concentration in the solid phase. Each line represents the relationship between these concentrations for a single soil or sediment. A straight-line segment through the origin

an organic compound onto three different media. In this case, the compound has a higher affinity, i.e., higher ratio of solid phase concentration to water concentration, for sediment than the other two substrates. Data from X. Xia, Z. Dai, J. Zhang, Sorption of phthalate acid esters on black carbon from different sources, J. Environ. Monit. 13(10) (2011) 2858–2864.

Concentration in solid medium (µg kg–1)

FIG. 13.13 Example sorption isotherms for 20

16

12

8

4

0

5

10 15 20 25 Concentration in water (µg L–1)

2. WASTE STREAMS (AND THEIR TREATMENT)

30

35

285

8 WASTEWATER MECHANISMS

represents the data well for the range of concentrations shown. Not all portions of an isotherm are linear and this is particularly true at high concentrations of the contaminant. Linear chemical partitioning can be expressed as: S ¼ KD  CW

(13.24)

where S ¼ concentration of contaminant in the solid phase (mass of solute per mass of soil or sediment), CW ¼ concentration of contaminant in the liquid phase (mass of solute per volume of pore water), and KD ¼ partition coefficient (volume of pore water per mass of soil or sediment) for this contaminant in this soil or sediment. For many chemicals and the substrates (e.g., soils), the partition coefficient can be estimated using: KD ¼ KOC  Corg

(13.25)

Concentration on solid surface (Csorb)

where KOC ¼ organic carbon partition coefficient (volume of pore water per mass of organic carbon) and Corg ¼ substrate organic matter (mass of organic carbon per mass of substrate, e.g., soil). KOC values are published in handbooks and manuals. However, it is important to note that these, like other partitioning coefficients, are substrate dependent. Thus when comparing

n>1

n=1

values, the values must be for the same substrate (e.g., carbon source ¼ anthracite) or otherwise normalized. The relationship between KD and KOC helps to approximate the “data poor” information from “data rich” information, that is, a chemical’s unknown KD from the known KOC of the contaminant and the organic carbon content of the soil horizon of interest. The actual derivation of KD is: KD ¼ CS ðCW Þ1

(13.26)

where CS is the equilibrium concentration of the solute in the solid phase and CW is the equilibrium concentration of the solute in the water. Therefore the solid waste manager may find this relationship useful in linking solid matrices with water, given that KD is a soil parameter. Indeed, KD is a direct expression of the partitioning between the aqueous and solid (soil or sediment) phases. A strongly sorbed chemical like a dioxin or the banned pesticide DDT in a landfill or waste site, for example, can have a KD value exceeding 106. Conversely, a highly hydrophilic, miscible substance like ethanol, acetone, or vinyl chloride, will have KD values <1. This relationship between the two phases demonstrated by Eq. 13.27 and Fig. 13.14 is the Freundlich Sorption Isotherm: FIG. 13.14 Hypothetical Freundlich isotherms with exponents (n) less than, equal to, and greater than 1, as applied to the equation Csorb ¼ KFCn. Sources: R. Schwarzenbach, P. Gschwend and D. Imboden (1993). Environmental Organic Chemistry. John Wiley & Sons, Inc., New York, NY; and H.F. Hemond and E.J. Fechner-Levy (2000). Chemical Fate and Transport in the Environment. Academic Press, San Diego, CA.

n<1

Concentration in Water (Cw)

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Csorb ¼ KF  Cn

(13.27)

Increasing the organic matter content is elevated in substrates, that is, soil and sediment, is accompanied by an increase in the amount of a contaminant that is sorbed, that is, sorption is directly proportional to the soil/sediment organic matter content. This allows for a conversion of KD values from those that depend on specific soil or sediment conditions to those that are soil/sediment independent sorption constants, KOC:

where Csorb is the concentration of the sorbed contaminant, that is, the mass sorbed at equilibrium per mass of sorbent; KF is the Freundlich isotherm constant; and n is the exponent of the chemical concentration curve of the function of Csorb and CW in Fig. 13.14. The exponent determines the linearity or order of the reaction. Thus if n ¼ 1, then the isotherm is linear; meaning the more of the contaminant in solution, the more would be expected to be sorbed to surfaces. For values of n < 1, the amount of sorption is in smaller proportion to the amount of solution and, conversely, for values of n > 1, a greater proportion of sorption occurs with less contaminant in solution. Also note that if n ¼ 1, then Eq. 13.27 and the Freundlich sorption isotherm are identical [33].

KOC ¼ KD  ðfOC Þ1

(13.28)

where fOC is the dimensionless weight fraction of organic carbon in the soil or sediment. The KOC and KD have units of mass per volume. Table 13.2 provides the log KOC values that are calculated from chemical structure and those measured empirically for several organic

TABLE 13.2 Calculated and Experimental Organic Carbon Coefficients (KOC) Compared to Octanol–Water Coefficients (Kow) for Selected Contaminants Found at Hazardous Waste Sites Calculated

Measured

Chemical

log Kow

log Koc

Koc

log Koc

Koc (geomean)

Benzene

2.13

1.77

59

1.79

61.7

Bromoform

2.35

1.94

87

2.10

126

Carbon tetrachloride

2.73

2.24

174

2.18

152

Chlorobenzene

2.86

2.34

219

2.35

224

Chloroform

1.92

1.60

40

1.72

52.5

Dichlorobenzene,1,2-(o)

3.43

2.79

617

2.58

379

Dichlorobenzene,1,4-( p)

3.42

2.79

617

2.79

616

1,1-Dichloroethane

1.79

1.50

32

1.73

53.4

1,2-Dichloroethane

1.47

1.24

17

1.58

38.0

1,1-Dichloroethylene

2.13

1.77

59

1.81

65

trans-1,2-Dichloroethylene

2.07

1.72

52

1.58

38

1,2-Dichloropropane

1.97

1.64

44

1.67

47.0

Dieldrin

5.37

4.33

21,380

4.41

25,546

Endosulfan

4.10

3.33

2138

3.31

2040

2. WASTE STREAMS (AND THEIR TREATMENT)

287

8 WASTEWATER MECHANISMS

TABLE 13.2 Calculated and Experimental Organic Carbon Coefficients (KOC) Compared to Octanol–Water Coefficients (Kow) for Selected Contaminants Found at Hazardous Waste Sites—cont’d Calculated

Measured

Chemical

log Kow

log Koc

Koc

log Koc

Koc (geomean)

Endrin

5.06

4.09

12,303

4.03

10,811

Ethyl benzene

3.14

2.56

363

2.31

204

Hexachlorobenzene

5.89

4.74

54,954

4.90

80,000

Methyl bromide

1.19

1.02

10

0.95

9.0

Methyl chloride

0.91

0.80

6

0.78

6.0

Methylene chloride

1.25

1.07

12

1.00

10

Pentachlorobenzene

5.26

4.24

17,378

4.51

32,148

1,1,2,2-Tetrachloroethane

2.39

1.97

93

1.90

79.0

Tetrachloroethylene

2.67

2.19

155

2.42

265

Toluene

2.75

2.26

182

2.15

140

1,2,4-Trichlorobenzene

4.01

3.25

1778

3.22

1659

1,1,1-Trichloroethane

2.48

2.04

110

2.13

135

1,1,2-Trichloroethane

2.05

1.70

50

1.88

75.0

Trichloroethylene

2.71

2.22

166

1.97

94.3

o-Xylene

3.13

2.56

363

2.38

241

m-Xylene

3.20

2.61

407

2.29

196

p-Xylene

3.17

2.59

389

2.49

311

USEPA, (1996). EPA/540/R-96/018, Soil Screening Guidance: User’s Guide, Washington, D.C.: U.S. Environmental Protection Agency; USEPA, (1996). EPA/540/R-95/128 PB96-963502, Soil Screening Guidance: Technical Background Document, Washington, D.C: U.S. Environmental Protection Agency.

compounds and compares them to the respective Kow values.

8.1 Monod Equation In previous sections, we have considered metabolism of a single organism. Obviously, treating wastes depends on sufficient bacterial growth to degrade large amounts of organic matter. The Monod equation is an empirically derived expression of the rate of microbial biomass: μ¼

μmax S KS + S

(13.29)

where μ ¼ the specific growth rate of the microbe, μmax ¼ maximum specific growth rate, and Ks ¼ the Monod growth rate coefficient representing the substrate concentration at which the growth rate is half the maximum rate. The μmax is reached at the higher ranges of substrate concentrations. Ks is an expression of the affinity of the microbe for a nutrient, that is, as Ks decreases the more affinity that microbe has for that nutrient (as expressed by the concomitantly increasing μ). Thus wastewater treatment must bring the waste into contact with the microbes and enhance the degradation environment, such as by activated sludge. Engineered

2. WASTE STREAMS (AND THEIR TREATMENT)

288

13. WASTEWATER

systems may also add activated microbes (if indigenous microbes are not already breaking down the chemical). As these conditions change, the microbes undergo a series of stages [12]: 1. Lag phase: Upon initial exposure of the microbes to the chemical contaminant, sufficient time is needed for the organisms to become acclimated. 2. Accelerated growth rate phase: Following acclimation, the microbes propagate at an increasing rate. There are two major processes that allow for the degradation of chemical compounds by microbes: (A) The most effective treatment occurs when the organisms use the contaminant exclusively as a food source for their growth, metabolism, and reproduction. As mentioned, this is accomplished by the microbe’s ability to produce enzymes that catalyze the chemical contaminant as a carbon source. Recall that the chemical is the microbe’s oxygen acceptor during respiration. The microbe produces enzymes that hasten the process of breaking chemical bonds and transferring electrons from the contaminant to an electron acceptor (oxygen for aerobic respiration, and metals (e.g., iron and manganese) and inorganic chemicals (e.g., nitrates and sulfates) for anaerobic respiration). (B) A second, less effective process is known as “secondary utilization.” Microbes can transform chemical contaminants although the reaction provides no direct benefit to the microbial cell. Probably the most common, at least best-understood secondary utilization process is cometabolism, wherein the microbes break down chemicals coincidentally with enzymes that they normally synthesize for metabolism or detoxification. A case is the methane-oxidizing bacteria that happen to degrade chlorinated

3.

4.

5. 6. 7.

hydrocarbons, benzene, phenol, and toluene by producing enzymes needed to transfer electrons to methane (i.e., the bacterium’s normal electron acceptor). The normal methane oxidation enzymes auspiciously degrade the chemical contaminants, even though the chemicals cannot serve as the primary food source for the bacteria. Exponential phase: The cell mass and the number of cells are growing exponentially by binary fission. Declining growth phase: Cell mass and numbers of cells continue to grow, but at a decreasing rate. This is usually due to depletion of the food and electron source (i.e., the contaminant, hopefully). Other limiting factors could also come into play, such as some of the newly generated chemicals (“degradates”) inhibiting the growth of the microbes due to their toxicity. Stationary phase: Cell decay is about equal to cell propagation during this time. Decay: Cell decay now exceeds cell propagation. Exponential death phase: Cells are dying exponentially as cells no longer grow or propagate. Under optimal conditions, this indicates successful bioremediation because the food source (toxic organics) is entirely used up.

The cycle will repeat if the microbes are again introduced to another slug of food (organic compounds).

9 WASTEWATER DATA AND KNOWLEDGEBASES The waste manager needs reliable knowledgebases. The water knowledgebase must include key aspects of both water resources and water quality data. The U.S. government makes both types available. The U.S. Geological

2. WASTE STREAMS (AND THEIR TREATMENT)

REFERENCES

Survey (USGS) maintains excellent databases to support water resource decisions, including aquifer features like storage capacity and transmissivity. To a lesser extent, the USGS also gathers data on water quality. Conversely, the U.S. EPA has a greater focus on water quality, limiting its interest in water resources to potable water supplies and ground water contamination [34]. Often, this part of the knowledgebase is not owned by most users, but is taken from generally available databases, especially those housed on governmental and institutional portals and websites.

References [1] United States. Federal Water Pollution Control Act Amendments of 1972. Public Law. 92-500, Enacted on October 18, 1972. [2] D. Vallero, Translating Diverse Environmental Data Into Reliable Information: How to Coordinate Evidence from Different Sources, Academic Press, 2017. [3] M.E. Tiemann, Safe Drinking Water Act Amendments of 1996: overview of PL 104-182, in: Congressional Research Service, Library of Congress, 1999. [4] U.S. Environmental Protection Agency, Identification of test procedures, in: 40 CFR 136.3, Office of Water, Washington, DC, 2017. [5] U.S. Environmental Protection Agency, Clean Water Act Analytical Methods, Available from: https:// www.epa.gov/cwa-methods, 2018. [6] T. McKone, W. Riley, R. Maddalena, R. Rosenbaum, D. Vallero, Common issues in human and ecosystem exposure assessment: the significance of partitioning, kinetics, and uptake at biological exchange surfaces, Epidemiology 17 (6) (2006) S134. [7] R.P. Schwarzenbach, P.M. Gschwend, D.M. Imboden, Organic acids and bases: acidity constant and partitioning behavior, in: Environmental Organic Chemistry, John Wiley & Sons, Inc., New York, 1993, pp. 245–274 [8] N. R. Council, Use of reclaimed water and sludge in food crop production, National Academies Press, Washington, DC, 1996. [9] M. Johnson, Wastewater Laboratory Basics, in: presented at the IWEA Laboratory Workshop, Springfield, Illinois, 2013. Available from: https:// www.iweasite.org/past_conference_slides.php. [10] W. Telliard, Method 1684: Total, Fixed, and Volatile Solids in Water, Solids, and Biosolids, US Environmental Protection Agency, Washington, 2001.

289

[11] W. E. Federation and A. P. H. Association, Standard Methods for the Examination of Water and Wastewater, American Public Health Association (APHA), Washington, DC, 2005. [12] Delzer, G.C., McKenzie, S.W., Five-day biochemical oxygen demand: U.S. Geological Survey, Reston, Virginia; Techniques of Water-Resources Investigations, Book 9, Chapter A7, Section 7.0, 2003, http://pubs.water.usgs. gov/twri9A/; accessed on November 13, 2018. [13] C.P. Gerba, I.L. Pepper, Wastewater treatment and biosolids reuse, in: Environmental Microbiology, second ed., Elsevier Inc., New York, 2009, pp. 503–530. [14] D. Vallero, Environmental Biotechnology: A Biosystems Approach, Elsevier Science, 2015. [15] V. Novotny, Water Quality: Diffuse pollution and watershed management, John Wiley & Sons, New Jersey, 2003. [16] P.L. McCarty, Biotic and abiotic transformations of chlorinated solvents in ground water, in: Symposium on Natural Attenuation of Chlorinated Organics in Ground Water, Dallas/TX, 1996, pp. 5–9. [17] B.E. Rittmann, P.L. McCarty, Environmental Biotechnology: Principles and Applications, Tata McGraw-Hill Education, New Delhi, 2012. [18] L.P. Wackett, Co-metabolism: is the emperor wearing any clothes? Curr. Opin. Biotechnol. 7 (3) (1996) 321–325. [19] P.P. Egeghy, J.J. Quackenboss, S. Catlin, P.B. Ryan, Determinants of temporal variability in NHEXASMaryland environmental concentrations, exposures, and biomarkers, J. Expo. Sci. Environ. Epidemiol. 15 (5) (2005) 388–397. [20] D.A. Vallero, Fundamentals of Air Pollution, Fifth ed., Elsevier Academic Press, Waltham, MA, 2014, 999. [21] F. Baquero, J.-L. Martı´nez, R. Canto´n, Antibiotics and antibiotic resistance in water environments, Curr. Opin. Biotechnol. 19 (3) (2008) 260–265. [22] S.L. Gorbach, Antimicrobial use in animal feed—time to stop, Mass Medical Soc, N. Engl. J. Med. 345 (2001) 1202–1203. [23] V. Economou, P. Gousia, Agriculture and food animals as a source of antimicrobial-resistant bacteria, Infect. Drug Resist. 8 (2015) 49. [24] C. Walsh, Antibiotics: Actions, Origins, Resistance, American Society for Microbiology (ASM), Washington, DC, 2003. [25] M. La Farre, S. Perez, L. Kantiani, D. Barcelo´, Fate and toxicity of emerging pollutants, their metabolites and transformation products in the aquatic environment, TrAC, Trends Anal. Chem. 27 (11) (2008) 991–1007. [26] S. Weigel, U. Berger, E. Jensen, R. Kallenborn, H. Thoresen, H. H€ uhnerfuss, Determination of selected pharmaceuticals and caffeine in sewage and seawater from Tromsø/Norway with emphasis on ibuprofen and its metabolites, Chemosphere 56 (6) (2004) 583–592.

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[27] T. Vasskog, T. Anderssen, S. Pedersen-Bjergaard, R. Kallenborn, E. Jensen, Occurrence of selective serotonin reuptake inhibitors in sewage and receiving waters at Spitsbergen and in Norway, J. Chromatogr. A 1185 (2) (2008) 194–205. [28] D.P. Morgan, C.C. Roan, Absorption, storage, and metabolic conversion of ingested DDT and DDT metabolites in man, Arch. Environ. Health 22 (3) (1971) 301–308. [29] D. Vallero, Paradigms Lost: Learning From Environmental Mistakes, Mishaps and Misdeeds, Butterworth-Heinemann, Burlington, MA, 2005. [30] W. Lyman, Transport and transformation processes, in: Rand GM Fundamentals of Aquatic Toxicology. Effects, Environmental Fate, and Risk Assessment, Second ed., Taylor and Francis, Washington, DC, 1995, pp. 449–492. [31] D. Richter, in: D.A. Vallero (Ed.), Ion Exchange, Duke University, Durham, North Carolina, 1996. [32] J. Westfall, Adsorption mechanisms in aquatic surface chemistry, in: W. Stumm (Ed.), Aquatic Surface Chemistry: Chemical Processes at the Particle-Water Interface, vol. 87, John Wiley & Sons, New York, 1987. [33] D. A. Vallero, Environmental Contaminants: Assessment and Control. Amsterdam/Boston: Elsevier Academic Press, 2004, pp. xxxix, 801 p. [34] D.A. Vallero, Air pollution monitoring changes to accompany the transition from a control to a systems focus, Sustainability 8 (12) (2016) 1216.

[35] X. Xia, Z. Dai, J. Zhang, Sorption of phthalate acid esters on black carbon from different sources, J. Environ. Monit. 13 (10) (2011) 2858–2864.

Further Reading [36] J.N. Spencer, G.M. Bodner, L.H. Rickard, Chemistry: Structure and Dynamics, John Wiley & Sons, New York, 2010. [37] SMOC, (2003). The Sound Management of Chemicals (SMOC) Initiative of the Commission for Environmental Cooperation of North America: Overview and Update. Available from: http://www3.cec.org/ islandora/en/item/2031-overview-and-update-soundmanagement-chemicals-smoc-initiative-en.pdf. [38] D. Vallero, Environmental Contaminants: Assessment and Control, Academic Press, 2010. [40] USEPA, EPA/540/R-96/018, Soil Screening Guidance: User’s Guide, U.S. Environmental Protection Agency, Washington, DC, 1996. [41] USEPA, EPA/540/R-95/128 PB96-963502, Soil Screening Guidance: Technical Background Document, U.S. Environmental Protection Agency, Washington, DC, 1996.

2. WASTE STREAMS (AND THEIR TREATMENT)