September 18, 2013 EE&T Project No. 5327-02 Subject: Impact of Bromide Discharges into Drinking Water Sources from Coal-Fired Power Plants

BACKGROUND In December 2011, the US Environmental Pr otection Agency (USEPA) approved new strict discharge limits targeting oil- and coal-fired electrical pow er generating facilities. These requirements are scheduled to take effect in 2015, though MATS has been revised and “reconsidered”, and the comment period has b een extended, a number of times since first published (USEPA 2012 – USEPA approved in December 2011 but not published in Federal Register until February 2012). MATS will target reduc tions in emissions of metals (mercury, arsenic, chromium, nickel), acid gases (hydrochlo ric and hydrofluoric), part iculate matter, sulfur dioxide (SO 2 ), and nitrous oxides (NO x ). Figure 1 shows a USEPA prepared map of the US facilities with coal, oil, or bo th coal and oil units at the power plant (USEPA 2011). It is estimated that 1,100 coal-fired units and 300 oil-fire d units at 600 power plants will be impacted by the MATS requirements (USEPA 2011). In the remainder of this memorandum, the term “coal-fired” power plants will be us ed, but this will be interpreted to mean coal or oil-fired units at electrical power generating facilities with these units. This memorandum deals with how these new requirements to control emissions from coal-fired power plants may lead to discharges of more bromide into dr inking water sources, and how these bromide discharges will impact the pr oduction of compounds harm ful to public health and the ability of public water systems to m eet federal and state dr inking water treatment requirements. The discussion below will first discuss how some of the new MATS requirements may lead to more bromide discharges into drinki ng water sources, how this will impact drinking water treatment operations and resulting quality of drinking water delivered to customers, and some case studies illustrating impact of bromide discharges from coal-fired power plants on downstream drinking water systems.EE&T, IN Fi g ure 1

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dized d by ry has e ntly, h ighly o ntrolEE&T, INC. 3 devices. Addition of bromide-containing salts (usu ally calcium bromide) to the coal combustion unit can convert the mercury into the more water soluble Hg 2+ form. This soluble mercury is better removed by the wet scrubbers used to clean flue gasses. However, while the mercury in the wastewater can be removed prior to discharg e into a receiving stream, the added bromide is not well removed and typically ends up being di scharged. Some data regarding the fate of bromide in Dutch power plants equipped with se lective catalytic reduction (SCR), ESP, and wet scrubbers suggests that 82% of the total br omide added is discharged (Meij, 1999). In 14 full scale coal-fired power plant tests using calcium bromide (CaBr 2 ) to oxidize elemental mercury, greater than 90% of the mercury was oxidized with the addition of 25 to 300 ppm bromide by weight of coal (Chang et al., 2008). This range is wide due to the coal’s natural abundance of chlorine and bromine giving the co al varied natural performance for oxidizing mercury. A 1 MW power plant operated 24 hours a day, 365 days a year would produce 8,760,000 KWh/year. According to the US Energy Information Administration it takes 1.07 lbs of coal to produce 1 kWh. This means that fo r each MW of electrical power plant capacity, a coal-fired power plant would require 9,373,200 lbs coal/year or 25,680 lbs coal/day. With the range of 25 to 300 ppm for bromide noted in Ch ang et al. (2008), the amount of bromide added to the system each day per MW of power is estimated as follows:

25,680 ݈ܽ݋ܿ	݈ܾ ܹܯ ∙ ݕ݀ܽ ൈ ݎܤܽܥ	݈ܾ	25 ଶ 10 ଺ ݈ܽ݋ܿ	݈ܾ ൈ ݎܤ	݈݋݉	2 ݎܤܽܥ	݈݋݉	1 ଶ ൈ ݎܤܽܥ	݈݋݉	1 ଶ 199.9	݃ ൈ 79.9	݃ ݎܤ	݈݋݉	1 0.51 ൌ ݎܤ	݈ܾ ܹܯ ∙ ݕ݀ܽ

25,680 ݈ܽ݋ܿ	݈ܾ ܹܯ ∙ ݕ݀ܽ ൈ ݎܤܽܥ	݈ܾ	300 ଶ 10 ଺ ݈ܽ݋ܿ	݈ܾ ൈ ݎܤ	݈݋݉	2 ݎܤܽܥ	݈݋݉	1 ଶ ൈ ݎܤܽܥ	݈݋݉	1 ଶ 199.9	݃ ൈ 79.9	݃ ݎܤ	݈݋݉	1 6.2 ൌ ݎܤ	݈ܾ ܹܯ ∙ ݕ݀ܽ If 0.51 to 6.2 lb/day of bromide are added for a 1 MW power plant, and 82% of the bromide is discharged as scrubber wastewater, theoretically 0.42 to 5.1 lb s/day of bromide can be discharged into receiving streams from a 1 MW plant. In 2006 about 100 gigawatts (GW) of coal-fired power plant cap acity in the United States was equipped with flue gas desulfurizat ion (FGD) technology, 90 GW of which are wet scrubbers (Miller et al, 2006). Fi gure 2 is from a National Ener gy Technology Laboratory report, projecting the flue gas desulfur ization capacity to increase to 231 GW by year 2020. If 90 percent of the FGD processes in stalled continue to be wet scrubbers, then 208 GW of theEE&T, IN electrica l b etween found in b romide fired boi l

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n ated have S EPAEE&T, INC. 6 2006). These adverse health effects have mostly been attributed to disinfection by-products (“DBPs”), the term used to collectively describe the suite of products resulting from reactions of free chlorine and other oxidants with natural or ganic matter (NOM) and other substances in the water. In particular, the DBPs most commonly associated withthese observed adverse health effects have been halogen substituted organics, typically DBPs containing chlorine or bromine. At present, while there are a large variety of complicated halogen su bstituted DBP compounds believed to result from chlorina tion of drinking wa ter, the only currently regulated DBPs are TTHM (the sum of chloroform, bromodichl oromethane, dibromochloromethane, and bromoform) and HAA5 (mono-, di-, and tri-chloro acetic acid and mono- and di-bromo acetic acid). Chlorine added to drinking water reacts quick ly with reduced substances (e.g., reduced iron), as noted above, plus it al so reacts quickly with ammonia, organic compounds containing nitrogen, and the bromide ion (Br - ). Any chlorine not consumed by these reactions can produce disinfection, but can also produce DBPs if en ough DBP precursor material is present and if unreacted free chlorine residual remains in water for a long e nough time for the DBP formation reactions to occur. Most chlorine in drinking water is consum ed in oxidation reacti ons, including oxidation of organics (Jolley 1975). However, some added chlorine substitutes into organic compounds to produce a chlorine substituted DBP, and some of the added chlorine can transfer its oxidative/disinfecting power to another compound. In the latter case, chlori ne can: a) react with ammonia to produce chloramines (typically monochloramine or NH 2 Cl), b) react with nitrogen containing organics to produce organic chloramine s, and c) oxidize bromide to bromine. When free chlorine reacts with bromide to produce free bromine in water, the bromine reacts analogously to free chlorine in: i) oxidation of any reduced metals still pr esent, ii) disinfection, iii) reaction with ammoni a to produce bromamines 1 (typically dibromamine), and iv) formation of bromine substituted DBPs if DBP precursor material is present.

1 Chloramines can also react withbromide to produce bromaminesEE&T, INC. 7 Therefore, in water chlorine can react direc tly with organic DBP precursors, but can also react first with bromide, then the resulting br omine can react with organic DBP precursors to produce brominated organic DBPs 2 . This is important for a number of reasons: 1. Greater health risks attributed to brominated DBPs than chlorinated DBPs = Risks from cancer and other adverse human hea lth effects are generally thought to be greater for bromine substituted DBPs than with analogous DBPs containing chlorine instead of bromine (Cantor et al. 2010). For example, the concentration of species that gives one in a million (1/1,000,000) lifetime cancer risk for dibromochloromethane (ChBr 2 Cl) is 0.6 μg/L on a mass basis. For chloroform (CHCl 3 ) it is 6 μg/L. Therefore, on a mass basi s it appears that CHBr 2 Cl has 10 times greater lifetime cancer risk than CHCl 3 . However, when you take in to account the different in molecular weight (119.4 μg/μmol for CHCl 3 and 208.3 μg/μmol for CHBr 2 Cl), the difference on a molar basis is ev en higher (~17.4 times greater). 2. Regulatory limits for drinking water comp liance are mass based not molar based = Drinking water facilities are currently faced with two regulatory limits for halogen substituted DBPs: TTHM (sum of four compounds listed above) ≤ 0.080 mg/L and HAA5 (sum of five compounds listed above) ≤ 0.060 mg/L. Consequently, the regulatory limit is simply based on taki ng the mass concentration of each compound, without correcting for molar weight, and ad ding each numerical value for the four or five compounds involved. Therefore, a wa ter system with no bromide in background source water will be in compliance with TTHM limit if they have 60 μg/L of chloroform and no detectable brominated TH Ms. In this case, about 0.5 μmol/L of TTHMs were produced. However, if nothing else changes but enough bromide is added to produce bromodichloromethane inst ead of chloroform, then 0.5 μmol/L of TTHM will produce ~82 μg/L which could create compliance difficulties 3 . Therefore, a water system currently in compliance w ith federal and state requirements may no

2 Furthermore, although free chlorine cannot react w ith bromine in drinking wa ter to produce bromate, a stronger oxidant like ozone can reach w ith bromide in drinking water to pr oduce bromate, which is a regulated inorganic DBP in drinking water. 3 Compliance is based on annual average at each compliance location in a water system, as required in Stage 2 DBPR. Therefore, a single TTHM value above 0.080 mg/L (80 μg/L) will not mean the water system is out of compliance, as long as th e average TTHM for four samples at this location during a one year period is ≤ 0.080 mg/L.EE&T, INC. 8 longer be in compliance if bromide is adde d to the drinking water source in amounts sufficient to increase the amountof brominated DBPs produced. 3. Increased formation of brominated DBPs due to greater reactiv ity of bromide = The above note suggests that the presence of bromide during chlorination can increase the numerical value of the regu latory compliance value even if the same amount of organic precursor material is present and even if the same molar concentration of DBP is produced. However, since free bromine in drinking water reacts more quickly than free chlorine, more DBPs (on a molar basis) will be produ ced (the increased reactivity of bromine will mean more oxida tion of organics by br omine, as well as bromine substitution reactions). Therefore, releases of bromide into drinki ng water sources can potentially impact public health for consumers of treated drinking water, and can substantially complicate treatment at drinking water facilities in or der to meet regulatory requiremen ts. Water systems have typically already completed their studies to evaluate source water quality in order to develop DBP control strategies. New bromide releases from coal-fired power plants ha ve not been evaluated by water plants beforehand. If these bromide releases occu r, the drinking water systems may not detect the changing conditions until they start noting elevated DBP levels, particularly brominated species, in compliance monitoring locations. Then all water systems in the watershed will have to do an immediate re-evaluation of treatment to meet the regulatory limits under the new conditions. Unfortunately, there is little the water system s can do once the bromide gets into the source water. They could try drasticallyincreasing removal of organic DB P precursors, but that will not be easy or affordable. Removing the bromide once it is in the water will be even more problematic. The best solution would be to find a way to keep the bromide from getting into the drinking water sources, either by: a) not using bromide to remove mercury at coal-fired power plants, b) using coal with lower mercury cont ent, or c) trying to remove bromide from wastewater streams prior to discharge (this will not be easy either, but will at least be easier to do in a concentrated stream than in a more dilute stream at the water plant).EE&T, INC. 9 CASE STUDIES While many US water systems do not have sour ce water bromide levels as high as some overseas locations, there has become a new concern associated with bromide discharges into potable water sources. There is some information available in the literature where drinking water facilities have detected an increa se in brominated DBPs and they have traced bromide discharges from coal-fired power plants as potential sources (States et al. 2013 and Hopkins 2013). These references also cite other potential bromide s ources, including hydrofrack ing, treated wastewater effluents, and indus trial discharges. Figure 1 indicated the USEPA identified coal- and oil-fired electrical generating facilities in the US. Except for a large numbe r of generating facilities in TX, most of the power plants indentified are east of the Miss issippi River, especially in TN, GA, AL, SC, NC, VA, eastern OH, western PA, and several locations along the Ohio River (IL, IN, OH, KY, and WV). Locations with fossil fuel powered electrical generating faciliti es in US generally use coal,, though oil-fired facilities can be found at some locations along either Atlantic or Pacific coasts (CA, NY, NJ, MA, CT, FL) and in some im portant oil producing areas (TX, OK, LA). Figure 4 includes locations of three power plants, rivers, streams, and a drinking water treatment facility using surface wa ter located upstream of Charleston SC. This situation is an example of a water system directly impacted by bromide discharges from a power plant. As shown in Figure 4, the City of Charleston surfa ce water treatment plant, called Hanahan Water Treatment Plant, uses the Cooper River as their wa ter source and their intake is downstream from three coal fired power plants. The power plant th at is furthest upstream has a capacity of 2,390 MW and is located at the north s hore of Lake Moultrie. The next power plant south of this power plant has a capacity of 346 MW and is located on the south shore of Lake Moultrie. The power plant closest to Charleston’s intake has a capacity of 633 MW and installed a wet scrubber in 2009. The power plant’s website indicates that th e new wet scrubber eliminates 60-90% of the former mercury emissions.EE&T, IN Figure 4 A which h a located i during J a occasion were as h

C. 4 . Three co a treatme n A s shown in

a ve been r o i n this wate r a nuary 200 2 ally were r a h igh as 1.9 m a l-fired pow n t plant for d

Figure 4, t o utinely mo n r shed, own e 2 thru July arely above 0 m g/L on two er plants i n d rinkin g w a there are so m n itore d for e d and sam p 2006 revea 0 .6 mg/L, b u occasions. n South Ca r a ter m e upstream bromide. T w p led by a S led that br o u t there we r r olina and a

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10 w ater rshed n s are t oring m g/L, e sultsEE&T, INC. 11 Figure 5 shows bromide levels at the finished water tap (blue line) and in the source water reservoir (red line) from 8/1/2008 to 9/13/2013. For the most part the bromide level from August 2008 to August 2010 stays around 0.1 mg/L and only exceeded 0.15 mg/L on a couple of days during this period. After August 2010 the bromide concentrations rise and peak around 0.35 mg/L in December 2010. Another rise occurred in October 2011, this time with peaks at 0.45 and 0.55 mg/L in March and June 2012, respectively. The increased bromide in the plant’s source water appears to occur shortly after the wet sc rubber installation at the power plant in 2009. Figure 6 shows concentrations (μg/L) of three specific DBPs, bromoform (blue), bromodichloromethane (red), and bromochloroacet ic acid (green) analyzed from the WTPs finished water tap from 8/1/2008 to 9/13/2013 (note DBP formation wo uld be even greater at in distribution system due to longer reaction time for DBP formation). The constituent impacted most during this period was bromoform. Fr om August 2008 to June 2010 bromoform had a concentration around 1 μg/L but clearly increases by a factor of 10 to 20, r eaching levels of 17 to 25 μg/L Bromodichloromethane had peaks at 13 and 15 μg/L in June 2009 and August 2009. After that, the concentrations stayed below 5 μg /L until early 2013 when concentrations began to increase again to 10 μg/L. Bromochloroacetic acid concentrations remained consistently below 10 μg/L except for one spike at 40 μg/L on March 2013. Many other water treatment plants in South Carolina made similar observations, howev er they declined to provide these data. The water system evaluated whether drought/climate change could have caused the increased bromide levels but drought was ruled out as a direct or indirect contributor to the observed bromide and DBP results.EE&T, IN Fi g ure 4

Fi g ure

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Acid aterEE&T, INC. 13 SUMMARY New requirements for increased SO 2 removal at coal-fired po wer plants should increase releases of bromide discharges drinking water so urces. Because of the mercury content of some coal sources used at US coal-fired power pl ants, and new requirements to improve mercury removal, coal-fired power plants may increas e the use of bromide to oxidize and remove mercury in wet scrubber processes. However, adding the bromide to increase mercury removal will further increase bromide releases in wastewat er discharges from these facilities. This can have substantial impacts on the amount and t ype of DBPs produced during drinking water treatment at downstream facilities, particularly the increased production of brominated DBPs. These brominated DBPs have potential for greater health effects for drinking water consumers and may cause drinking water facili ties currently in compliance with DBP requirements to be out of compliance if a greate r proportion of the DBPs pr esent are brominated. Removing the bromide at the drinking wate r plant is not econom ically sound. Water systems are left without options to reduce DBPs and comply with USEPA drinking water regulations, absent installing co mbinations of drinking water treatment unit op erations (e.g., reverse osmosis (RO), granular activated carbon in combination with split-stream RO, etc.). Installing such advanced treatment will dramatica lly increase treatment costs and water rates in the communities served, as well as potentially creating other unintende d consequences (e.g., disposal of membrane residuals, disposal or re generation of spent carb on, disposal of anionic exchange regeneration brine wast es, etc.). Removing bromide at the discharge source is less expensive and more proactive than to remove diluted bromide at multiple downstream water treatment facilities. It is also more equitable for the discharger to remove the bromide rather than forcing downstream water users to have to addre ss this precursor to cancer causing contaminants. REFERENCES AWWA 2004. Formal Comments of the American Wate r Works Association on the Stage 2 Disinfectants and Disinfection By-Produc ts Proposed Rule, Water Docket OW-2002- 0043 . Washington, DC: AWWA. Blythe, G.; Richardson, C.; Rhudy, R. 2002. Pilot Evaluation of the Catalytic Oxidation of Mercury for enhanced Removal in Wet FGD Systems. In Proceedings of Air Quality III: Mercury, Trace Elements and Particulate Matter Conferen ce, Arlington, VA, September 9-12, 2002. Energy and Environmental Research Center, University of Dakota, Grand Fork, North Dakota, September 2002.EE&T, INC. 14 Bustard, J.; Renninger, S.; Chang, R.; Miller, R.; Monroe, L.; Sjostrom, S. 2003. Results of Activated Carbon Injection for Mercury Control Upstream of COHP AC Fabric Filter. Presented at the A&WMA/EPA/DOE/EPRI Combin ed Power Plant Air Pollutant Control Mega Symposium, Washington, DC, May 19-22, 2003

. Chang, R. et al. 2008. Near and Long Term Opti ons for Controlling Me rcury Emissions from Power Plants. Proceedings of the Power Plant Air Pollutant Control “Mega” Symposium, Baltimore, MD, August 25-28, 2008. Hopkins, D. 2013. Bromide – an unsuspected culprit assessing impacts on Stage 2 DBPR compliance. Presented at VA-AWWA/VWEA Joint Annual Meeting, September 8-12, 2013. Richmond, VA. Jolley, R. 1975. Chlorine-containing organic constituents in chlorinated effluents. Jour WPCF

. 47:3:601. Kellie, S.; Cao, Y.; Duan, Y.; Li, L.; Chu, P.; Meht a, A.; Carty, R.; Riley, J.T.; Pan, W.P., 2005. Factors Affecting Mercury Speciation in a 100-MW Coal-Fired Boiler with Low-NOx Burners. Energy Fuels , 19:3:800-806. Mcilvaine Company, 2012. Big Market fo r Chemicals to Capture Mercury. www.pollutionsolutions-online.com Meij, R., 1999. Mass Balance Study.... presented at IEA Trace Element Workshop, University of Warwick. Miller, C., T. Feeley, W. Aljoe, B. Lani, K. Schr oeder, C. Kairies, A. McNemar, A. Jones, J. Murphy. 2006. Mercury Capture and Fate Usin g Wet FGD at Coal-Fired Power Plants. DOE/NETL Mercury and Wet FGD R&D . August 2006. States, S., et al. 2013. Marcellus Shale drilling an d brominated THMs in Pittsburgh, PA, drinking water. Jour AWWA . 105:8:E432-E448. US Energy Information Administration. 2013. How Much Coal, Natural Gas, or Petroleum Is Used to Generate a Kilowatt-hour of Electricity? Web. <http://www.eia.gov/tools/faqs/faq.cfm?id=667&t=2>. USEPA. 2006. Stage 2 Disinfectants and Disinfection Byproducts Rule: National Primary and Secondary Drinking Water Regulations: Final Rule. Federal Register 71:2:388-493, January 4, 2006. USEPA. 2011. Regulatory Impact Analysis for the Fi nal Mercury and Air Toxics Standards (EPA-452/R-11-011). Washington, DC: USEPA. USEPA. 2012. National Emissions Standards for Hazardous Air Pollutants from Coal- and Oilfired Electrical Generati ng Units (40 CFR 60 and 63). Federal Register . 77:32:9304- 9513. White, G. 1986. Handbook of Chlorination, 2nd ed. , New York: Van Nostrand Reinhold.