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University of Alberta

T h e use of constructed wetlands to treat oi1 sands wastewatet-. Fort YcYurray. Alberta, Canada.

Farida Susanne Bishay

.A thesis submitted to the Faculty of Craduate Studies and Researcli in partial Tulfillment of the requirements for the degree of 'laster of Science.

i r i

Env ironmen ta1 Biolog y a11 d Ecdogy

Department of Biologicai Sciences

Edmonton. Alberta

Spring 1998

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Dedicated To

Daddy. Assaad Bishay ( 1935- 1977) Yormor, Asta Schaltz ( 1898- 1993)

.Auntie Sadia, Sadia Bishay ( 1927- 19%) Cncle Samy, Samy Bishay (1935-1996)

They taught me that to run is to discover,

to rernember is t u unders tand, to teach is to learn,

and to write is to share.

ABSTRACT

Constructed wetIands are a component of the mine reclamaliori plan for-

Suncor Inc., OiI Sands Croup. a large oil sands mining and extraction

operation in northeastern Alberta, Canada. Over a three year period (1992

- 1994), the feasibility of constructed wetlands as treatrnent systems for

various oil s a n d s w a s t e ~ a t e r streams was assessed. T h e removal rates of

ammonia and hydrocarbon, t h e primary contaminants. were estimated b ~ -

rneasuring inputs and outputs , changes in storage. and removal processes.

Removal ra tes ranged from 32 to 99 % and 19 to 76 % for ammonia and

hyd rocar bons. respectiveiy. Removd ra tes decreased a s in pu t load

increased. Loads removed w e r e about 200 mg rn-' d-' and 170 mg m.' d" for

ammonia and hydrocar bons, respectively (on average during the period of

wastewater application ). Yeasurements of s torage and transformation

processes suggested t h e dominant fate path ways ciere sed imenl retention

and nitr if ication/denitrification for ammonia and sedimen t re ten tion and

microbial mineralization for hydrocarbons. Yacrophyte production and

der'omposition were rneasured tu assess viability and sustainabilitj-.

PI-oduction (as measured b y peak s t a n d i n g crop) alid decomposition (as

measured by weight loss) were only significantlj- different among

treatments in 1993 under higher hydraulic loading ra tes of 4.9 cmid

compared with 1.6 cm/d in 1992 and 1994. Production [meari 2 SD) in 1993 1

was 585 + 243. 470 2 175 and 453 i 216 g,/rnm in Control. D y k e and Pond

trenches. respectively. Decomposition (mean I S D ) in 1993 w a s 52.2 2 9.1

%. 47.2 I 4.2 % and 35.6 -+ 4.9 % in Control, Dyke and Pond trenches.

respectively. The treatment wetlands provided significant contaminant

retention under l ighter loading. and macrophyte production and

decomposition were not impacted a t these lighter loadings. However.

longer-term. larger-scale demonstrations would be required to determine

via bility and sustaina bility of constructed wetlands. Such facilities a r e

currently planned and are considered the next s t e p by the oil s ands

industry before design and implernentation of ful l scale constructed

wetlands.

This research was a cooperative e f for t between Suncor Inc. - Oil Sands Group of Fort McYurray, AB (Suncor): EVS Consultants L td of \o r th Vancouver. BC (EVS); and Lnivers i ty of Alberta. Edmonton. AB. This research kas funded by Suncor. Y r . John Gulley (Suncor) . >Ir. Peter \ix (EVS) and Yr. Steve Tut t le (Suncor ) provided financial. logistical and moral s u p p o r t th roughout this project.

I would like to thank m y superv isor , D r - Suzanne Bayle)., for he r expeditious review, enthusiasm, sage advice a n d moral suppor t du r ing th i s under taking. In addition. 1 would like to thank t h e o the r c o r n m i t t e e rnembers, Drs. Bill 'IcGi11 and Yichael Hickrnan. for their input and advice.

T h i s research project afforded m e t h e opportunity to worL with man3 people affiiiated with Suncor, EVS and t h e Lniversity of Alberta. 1 greatly

apprecia ted the i r patience and unders tanding when t h e project w a s running rough and their smiies when i t was running smoothly. I would especially Iike to thank Ys. Trina Hoffarth and .Ir. L m Paquin of Suncor fo r the i r hard work and never ending senses of humour. In addition, 1 would l ike to thank everyone from Suncor 's Chemistry Laboratory, and especially >Ir. Terry Arbic, 'Ir. Bob Benko, Yr. Ron l lart in and 'Ir. Bill Karning who were always helpful in rny greatest times of need regard iess of how expensive t h e piece of equipment I wanted to borroh. kas. I would

a lso iike to thank Lis. Gerti Hutchinson and 11s. Patricia Burgess of t he University of Alberta's Limnology Laboratory for analyzing m y seerningly

infinite number of sampies. 1 would like to thank D r . YcGi l l for providing access to gas chromatographie equipment and Y r . Reynald Lemke for his time and patience in ins t ruct ing m e on the use of the equ ipment for t h e analysis of N,O. Statistical advice w a s gladly proc-ided by D r . Yichael Paine of pain; Ledge and Associates. Vancouver. BC. Dr. Hugh Hamilton of Summit Environmental Consulting and Dr. Leah Bendell Young of Simon F ra se r Cniversi ty also provided moral a n d technical suppor t and advice.

And finally 1 would like to thank my family and friends. \N-ho always managed t o rnotivate me over t h e telephone iines ... Thank you! Especiaily. Yom and Bill fo r giving me a mmforting place to corne and hide when 1

needed a break: the Weibe family and Dee Ann for giving m e a place t o stay in Edmonton and putting u p with m y strange hours: and most especially to my sister. Soraya. who took care of every th ing else during th i s undertaking.

TABLE OF CONTENTS

Page No . . . . . . . . . . . . . . . . . . . . . . . . . . . FNTRODUCTION 1

Sunmr and theoil sands extraction process . . . . . . . 1 Wetlands as wastewater treatment facilities . . . . . . . . 3

. . . . . . . . . . . . . . . . . . . Objectives and Approach 4 . . . . . . . . . . . . . . . . . . . . . . . . . Literature Cit& I l

AMMONIA REMOVAL IN CONSTRUCTED WETWDS . . . . . . . . . . . . . RECEMNG OIL SANDS WASïZWATER 13

. . . . . . . . . . . . . . . . . . . . . . . . . . . Introduction 13 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Methods 16

. . . . . . . . . . Site Description and Experimental Design 16 . . . . . . . . . . . . . . . . . . . . . . . . Sample Coilection 18

. . . . . . . . . . . . . . . . . . . . . . . . . Sample Analysis 18 . . . . . . . . . . . . . . . . . . . . . Water Chemistry 18

. . . . . . . . . . . Sediment and Tissue Chemistry 19 . . . . . . . . . . . . . . . . . . . . . In situ Denitrification 19 . . . . . . . . . . . . . . . . . . . . . Ammonia V o l a m t i o n 20

. . . . . . . . . . . . . . . . . . . . . . . . . Nitrogen Budget 21 . . . . . . . . . . . . . . . . . . . . . . . Statisticai Analysis 22

. . . . . . . . . . . . . . . . . . . . . Water Chemistry 22 . . . . . . . . . . . Sediment and Tissue Chemistry 22 . . . . . . . . . . . Volatilj7;ition and Denitrification 23

. . . . . . . . . . . . . . . . . . . . . Results and Discussion 23 . . . . . . . . . . . . . . . . . . . . . . . . . Water Chemistry 23

. . . . . . . . . . . . . . . . . . . . . . . Sediment Chemistry 23 . . . . . . . . . . . . . . . . Macrophyte Tissue Chemistry 27

. . . . . . . . . . . . . . . . . . . . . . . . . . Denitrif ication 28 . . . . . . . . . . . . . . . . . . . . . . . . . . . Voiatili7;ition 30

. . . . . . . . . . . . . . . . . . . . . Nitrogen Mass Balance 31 Treatment Efficiency . . . . . . . . . . . . . . . . . . . . . . 33

. . . . . . . . . . . . . . . . . . . . . . Long-term Treatment 36 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

. . . . . . . . . . . . . . . . . . . . . . . . . Literature Cited 32

HYDROCARBON REMOVAL IN CONSTRUCTED WETLANDS . . . . . . . . . . . . . RECEIVING OIL SANDS WASTEWATER 57 .. . . . . . . . . . . . . . . . . . . . . . . . . . . . Introduction a I

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Methoàs 60 . . . . . . . . . . Site Description and Experimental Design 60

. . . . . . . . . . . . . . . . . . . . . . . . Sample CoUection 62 . . . . . . . . . . . . . . . . . . . . . . . . . Sample Analysis 62

. . . . . . . . . . . . . . . . . . . . . . . . . . . TEH Budget 62 . . . . . . . . . . . . . . . . . . . . . . . Statisticai Anaiysis 63

. . . . . . . . . . . . . . . . . . . . . Water Chemistry 63 . . . . . . . . . . . . . . . . . . . Sediment Chemistry 63

. . . . . . . . . . . . . . . . . . . . . Re~ults and Discussion 64 . . . . . . . . . . . . . . . . . . . . . . . . . Water Chemistry 64

. . . . . . . . . . . . . . . . . . . . . . . . . . . TEH Removal 64 . . . . . . . . . . . . . . . . . . . . . . . Sediment Chemistry 65

. . . . . . . . . . . . . . . . . . . . . . . . TEH Mass Balance 65 . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions 66

. . . . . . . . . . . . . . . . . . . . . . . . . Literature Cited 74

EMERGENT AQUATIC MACROPHYTE PRODUCTION AND DECOMPOSITION IN CONSTRUCTEO WETLANDS RECEIVING

. . . . . . . . . . . . . . . . . . . OIL SANDS WASTEWATER Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Site Description and Experirnental Design . . . . . . . . . Production . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . . . . . . . . . . . . . . . . . Decomposition . . . . . . . . . . . . . . . . . . . . . . Statistid Analysis

. . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . Production

. . . . . . . . . . . . . . . . . . . . . . . . . Decomposition Condusions . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . . . . . . . . . . . . . . . . Literature Citeà

. . . . . . . . . . . . . . . . . . . . . . . . . CONCLUSIONS . . . . . . . . . . . . . . . . . . . . . Scope of the pmblem

The potential for the use of wetlands to treat oil sands wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . Do wetiands treat oil sands wastewater? How w e i l do wetlands treat this wastewater? . . . . . . Where do the mntaminants go? . . . . . . . . . . . . . . . What factors affect wetland treatment efficiency of oil sands wastewater and can these factors be optimked? 1s treatrnent sustainable? . . . . . . . . . . . . . . . . . .

. . . . . . . . . . AppLicabiiity to the Oii Sands Industry . . . . . . . . . . . . . . . . . . . . . . . . Literature Cited

APPESDIX .4 . Study Site Descriptions APPESDIX B . kater Chemistry Yethods and QA/QC APPEXDIX C . Hydrometeorology APPENIDX D . Nodel

Table i.1

Table 2.1

Table 2.2

Table 2.3

Table 2.4

Table 2.5

Table 2.6

Table 3.1

Table 3.2

Table 3.3

Table 3.4

Table 3.5

Table 4.1 Table 4.2 Table 4.3

Table 4.4

Table 4.5

Table 4.6

WST OF TABLES

Page No.

Physical, chernical and biological con taminan t removai rnechanisms in aquatic systems. Experimental design for t h e Constructed tc'etlands study. 1992- 1994. Ammonia-. nitrite-nitrate- a n d total k jeldah 1 nitrogen concentrations (mean) for t h e inFlow water ( A ) a n d oufflow water ( D ) of t h e constructed wetlands trenches. Ammonia removal in constructed wetlands receiving oil s ands wastewater. Sutr ient sedirnent chernistry (mg kg-') in the constructed wetlands t renches, Septem ber. Concentration of nutr ients in plant t i s s u e chemistry (mean) in t h e constructed wetlands trencves. Ammonia nitrogen mass balance ( kg trench- season-') for Dyke and Pond water. Total extractable hydrocarbon in water ( m g L-') for t h e inflow ( A ) and outflow ( D ) of t h e constructed wetlands trenches. Total extractable hydrocarbons (TEH) removal in constructed wetlands receiving oil s a n d s wastewater. Oil and g rease removal in various wetlands treatrnent facilities. Total extractable hydrocarbons in t h e sediments (mg kg' ) in t he constructed wetlands trenches.

Total Extractable Hyd rocar bon mass balance for Dy k e and Pond water. Experimental design for production measuremen ts. Experimental design for Iitter decomposition. Above-ground macrophyte biomass (mean 2 s tandard deviation) in t he constructed wetlands. 1992 - 1994. Below-ground macrophyte biomass (rnean + standard deviation) in t h e constructed wetlands. 1993. Decornposition (% reduction from initial mass. mean 2 S D ) of leaf l i t ter over various exposure periods from 1992 - 1994. Decay coefficients and half lives for litter under various water quality conditions.

Figure 1.1 Figure 1.2 Figure 1.3

Figure 2.1

Figure 2.2

Figure 2.3

Figure 2.4

Figure 2.5

Figure 2.6

Figure 2.7

Figure 2.8

Figure 2.9

Figure 2.10

Figure 3.1

Figure 3.2

Figure 4.1

Figure 4.2

Figure 4.3

LIST OF FIGURES

Page No.

Study s i tes around t h e Suncor mine site. Aerial photograph of Suncor Lease 86. Schematic of the Su ncor Constructed ke t lands researcfi facility. Arnmonia removal (%) for Dyke (1992 - 94) and Pond (1992 - 93) water. Nitrite-nitrate nitrogen concentration in the Dyke water replicate t renches at Stations A, B, C and D, 1994. Ammonia-ammonium nitrogen in the Dyke water replicate t renches at Stations A, B, C and D, 2994. Exchangeable amrnonia nitrogen (mean S D ) in t h e sediments of the constructed wethnds trenches, 1993 and 1994: by date. Exchangeabie ammonia nitrogen (mean S D ) in the sediments of t h e constructed wetlands trenches, 1993 and 1994: by zone. Total nitrogen in t h e storage pool in 1993 and 1994 at t h e end of the growing season (September). In situ denitrification (mean i S D ) measured us ing t h e acetylene blockage technique in Controt and Dyke Drainage trenches. 1994. Ammonia volatilization rates (mean S D ) From t h e Constructed Wetlands trenches, 1994. The nitrogen budget in t h e (a) Control and ( b ) Dyke t renches in 1994. Observed and predicted effluent ammonia concentrations (mean + SD), 1992 - 1994. A cornparison of CC chromatograms for a naphthenic acid s tandard and Pond inflow and outflow water. Hydrocarbon removal (%) for Dyke (1992 - 94) and Pond water (1992 - 93). Above-ground macrophyte biomass (mean + S D ) in the constructed wetlands, September 1993. Yass remaining ( d r y weight as % of initial mass: mean 2 SE) for Typha IafifoIia green leaves in 1993 iri the constructed wetland trenches. Mass remaining ( d r y weight as % of initial m a s : mean 2 SE) for T'pha latifolia green leaves in 1993 in t h e Nat ural Wetland and Reference Wetlands.

1.1 Sunmr and the oil sands extraction process

Suncor Inc., Oil Sands Group (Suncor) is the first of t w o cornpanies mining oil sands in northeastern Alberta, Canada, 23 km north of Fort YcYurray (Figure 1.1). The mining and subsequent conversion of oil sands to synthetic crude oil began in 1967. There are three elements to produce synthetic crude oil: (1) mining of the oil sands: (2) extraction of bitumen (oil) from the sand; (3) upgrading of the bitumen to synthetic crude oil. These three processes and related activities produce various deleterious wastes which are released directly or indirectiy (i.e., af ter some treatment) into the environment.

There are several environmental concerns associated with the production of synthetic crude oil. Yining practices directly alter vegetation patterns (by the removal of spruce and tamarack forests) and the hydrology of the region ( by draining land surfaces). Extraction of the bitumen produces fine tailings which are stored in several large tailings ponds (Figure 1.2).

The upgrading and power generation processes release pollutants (e-g.. SO,) t o the atmosphere.

While al1 these factors affect the environment, the cur rent s tudy ici11 focus on the problern of the accumulating fine tailings. In the p r e s s of extracting bitumen ( Le.. the Clark Hot hater Extraction Process) from the oil sands. a waste product called fine tails is produced. Fine tailings are an aqueous suspension of sand. silt. clay and residual bitumen and naphtha with a pH between 8 and 9 (FTFC, 1995a). Fine tailings have b e n stored in large ciay lined tailings ponds to date.

These ponds must be reclaimed to a viable land surface or water body free of long term maintenance (Nix and Yartin. 1992). There are two types of reclamation scenarios: dry landscape and w e t landscape (Gulley and MacKinnon. 1993). Dry landscape options include amending fine tails with soil o r spreading layers of fine tailings over large areas to undergo drying ( Le.. evaporation of associated water ) in summer and f reeze-thawing ( Le.. separation of water from clay by freezing and thawing) in winterhpring. The main objective of the d r y landscape recfamation plan is to significantly reduce the water content of the fine tailings such that a solid deposit is

produced. which could be capped with soil as part of' a terrestrial

reclamation plan ( FTFC, 1995 b ) . The wet landscape scenario deposits the fine tailings as a fluid requiring geotechnically secure containment. The objective of the w e t landscape reclamation plan is to isolate t h e f ine tailings under a layer of water which would act to esbbl i sh a viable. self- sustaining aquatic system above the f ine tailings layer (FTFC, 1993b).

There is a potential for water release from both d r y and wet landscape scenarios. -4lthough the water quality of these released waters is expected to be good (FTFC. 1995b), the contaminant ioad is unknown. To address this potential water release either a collection system must be maintained or a treatment system created. Wetiands would act as a polishing treatrnent system for water leaving t h e reclaimed site before entering the receiving environment (Gulley and Klym, 1992).

To meet t he demands of a reclamation plan, a wetland treatment system w a s investigated by Suncor. Two main types of water have b e n assessed in the experirnental constructed wetlands: leachate from saturated tailings pond sand dykes (as a surrogate for al1 leachate water types) and tailings pond top water (as a worst case scenario). Leachate water (from the tailings pond o r precipitation throughfall that has passed through sand dyke w a l l s and coke filters) is current ly gravi ty fed to a pumping station which r e tu rns it to the hil ings pond (or recycles it through the extraction plant). kater for the constructed wetlands was obtained frorn this pumping station and will b e referred to as Dyke Drainage or Dyke water. Tailings pond water (lower in suspended solids compared with fine tailings, bu t much higher compared with Dyke water) is decanted from one tailings pond (Pond 1) into another (Pond 1.4) containing recycle water only (i.e.. no fine tailings). Kater for the constructed wetlands was obtained frorn sur face waters of Pond 1A and will be referred to as Pond 1.4 or Pond water.

Organic (e-g.. naphthenic acids) and inorganic (eg., arnmonia) contaminants are of concern in both D y k e and Pond water. Although naphthenic acids exist naturally in bitumen (Fan, 1991) and are not formed during the extraction and refining process (AEP, 1996), t he re are elevated levels of solubiiized naphthenic acids in oil s ands fine tailings d u e to t he alkaline hot water extraction processing of oil s ands (.lacKinnon and Boerger, 1986). Naphthenic acids are present as sodium salts in oil sands wastewater (Herman et al., 1994). The primary source of ammonia is the Sour water. which contains both reduced sulphur and ammonia products. Sour water is the water that passes through the su lphur recovery and naphtha

recovery units. These recovery units clean u p the water by rernoving su lphur (and coincidentally ammonia) and recycling nap h tha. Essentially, wherever sulphur i s produced o r removed s o i s ammonia. The quality of these waste streams has improved a n d today much less ammonia is released into t h e tailings ponds via Sour water than previously. Secondary ammonia sources include the knock out pot (a collecter at the bottom of the flares) and s p e n t amine which is used as a su lphur scrubber .

These contaminants (Le., naphthenic acid and ammonia) a r e of concern as they are toxic to aquat ic organisms. A major source of acute toxicitÿ to rainbow t rou t has been traced to naphthenic acids extracted from oil s ands f ine tailings (AEP, 1996; MacKinnon and Boerger, 1986). Ammonia is also acutely toxic t o t rou t (USEPA. 1985). Cornparing resui ts from toxicity tests (Rainbow Trout 96 h LC5O) conducted by Thurston et al. (1981) with those conducted for Dyke water samples i t is apparen t that approximately 50% of t h e toxicity observed i s d u e t o ammonia (Bishay and Mx, 1996). Therefore a treatrnent system t h a t removes naphthenic acids and ammonia, and concomi tan tly the toxicity is req uired .

1.2 Wetlands as wastewater treatment facilities

het lands have been successfully used in treating domestic o r municipal wastes (e -g . , in nor thern Quebec b y Dubuc et al., 1986). urban runoff (e.g. , in Lake Tahoe Basin, Nevada by Reuter, 1992) and industrial wastes (e.g., acid mine drainage in Colorado by l o r e a et al., 1990). etc. A varie ty of physical. chernical and biologicai processes within wetlands are responsible for removing " target contaminants" (Table 1.1). There are only a Few examples (e-g., Amoco. Litchfield and Schatz, 1989: Imperia1 @il Ltd.. S i s et al., 1993) of wetlands being used to treat wastewaters from petrochemical facilities o r other sources of organic chemical wastes, and therefore Suncor has undertaken a s t u d y to a s ses s t h e treatment effectiveness of constructed wetlands on oil sands wastewater (Hamilton et al., 1993)-

A prelirninary s tudy of contaminated natural wetlands on t h e Suncor lease was conducted in October 1990 (Hamilton and Nlx, 1991). Both contaminant removal and toxicity reduction were observed from "inflow" to "outflow". For exarnple. total ammonia nitrogen was reduced 23% w hile total extractable hydrocarbons (TEH; which included naphthenic acids) were reduced 46%.

The c u r r e n t s tudy w a s undertaken based on these observations.

1.3 Objectives and A p p h

There are five major issues surrounding the potential u s e of wetlands. at a northern latitude. to treat oil s a n d s wastewater:

1. Do wetlands t reat oil s a n d s wastewater? - i-e,, are contaminant concentrations lower in the ou tflow ing water than inflowing water

2. How wel l do wetlands treat this wastewater? - Le., a t what r a t e are contaminants removed

3. hhere do the contaminants go? - i.e.. are they degraded, o r accumulated and where do these processes occur

4. hhat factors affect wetland treatment efficiency of oil sands wastewater and can these factors be optirnized?

- Le., are there minimum temperature requirements or maximum contaminant Loads

3. 1s treatment sustainable'? - Le., a r e wetland functions such as rnacrophyte productivity/decomposition o r nutrient cycfing affected by the wastewater o r treatment process such that effective treatrnent is not sustainable

To address these issues. nine constructed wetlands trenches were built in 1991 (Figure 1.3: Appendix A) . hastewater inputs began in 1992 and continued in 1993 and 1994. The flow rate of the wastewater varied year to year. Pond water w a s not tested in 1994. Instead, each of the three Pond water trenches were split in half lengthwise such that six flow rates of Dyke water couid b e tested. In addition to monitoring these constructed wetlands, the contarninated natural wetlands (Appendix A ) were also monitored to observe the potential for long term impacts. Reference wetlands (i.e., natural wetlands off t h e Suncor lease; Figure 1.1; Appendix A ) were monitored for cornparison between the treatment wetlands and uncontaminated natural wetlands.

Chapter 2 addresses i ) treatment efficiency by measuring ammonia inputs and outputs and estimating removal rates as t h e difference and by corn paring these observed removal rates w ith those predicted b y models developed in other studies. 2 ) contaminant (i.e., ammonia) fate by measuring nitrogen s torage pools ( i.e., sediment and macrophyte t i s sues) and physicai and biologicai removal processes ( Le. ammonia volatilization and nitrification/denitrif ication), 3) factors affecting treatment ef f iciency by monitoring removal ( e treatment efficiency) at various loads and considering the impact of temperature, pH, dissolved oxygen. nutr ients and hydrocarbons a s they pertain to ammonia removal.

Chapter 3 addresses 1) hydrocarbon removal efficiency by measuring hydrocarbon inputs and outputs to estimate removai by the difference, and 2 ) potential contaminant f a t e pathways were discussed.

Chapter 4 addresses treatment sustainability by assessing the potential impacts of the treatment waters on macrop h y t e productivity and decomposition tha t represen t wetland f unction.

Table 1.1 Ph ysicai. chernical and bioiogical contaminant removal mechanisms in aquatic systems. (Stowell et al. (1980).

reproduced by katson et al. (1989)).

9rchani sm ~ r f e c t ~ Cnntaminant Affectcd Drscrip~ ion

Phys ical

Scdimcntation Primary scltlcnblc solids Graviru scttiing solids rand

Secondaru col loidal sol ids constirurnt

contaminantsi in Incidcntal BOD. nitroucn. pond/marsh set t inas.

phos ph orus. hcavy

mctals. rcfractory

vi rus

Filtration Sccondary sc t L icnbir soi ids.

cnlloidai solids

Part iculatcs

l ' i l ~rrcd

mcrhonicaiIy as

w a l r r passes through

S U ~ S L ~ ~ L I - . rnot

masses. or f i s h .

.Adciorpi ion Sccondary col loidn! S O L ids Intcrparticic ai trncti\,c force it'an dcr \mals Corcc ) .

Prccipitn~ ion P r i m a r ~ . pnnsphnrris . ncavy Fnrmalion of' o r rnc:a;s coprccipi :at ion uith

i nsoluhlc compaunds .

.\dsorpt i o n rimary phosphnriis . hravy .\dsorpr ion on

m c ~ a l s substratc and plact

surfarc . Seconda ry rcfractory orgnnics

Dccomposition Primary rcfrac:tiry orgnnics Drcnmposi 1 ion or ni:rrat ion O C ~ P S S

slûblc compounds by

phcnomena such as I l '

irrndint ion.

osidalion and

rcdiic t ion.

Ycchanism ~ffccit" Contaminan1 .If f c c t c d Dcscript ion - -. - - - - - - -

Yicrobial Primary colloidal solids.

Yctabolism b BOD. nitrogcn. rrfracîot-y or~anirs

henry mrlais

Plant Sccondary rcf rac t ory ar~anics . 'ictabol ism b bactcria. virus

Sccondory ni t rogcn . phosphorus . hm\-y me tn?s . rcf ractory orRani C S

'ia t lirai Primary

Dicof t-

Rcmovai of rolloida:

s o i i d s and snlubic

orgon i rs hy

susprndcd . 5c~thir and plan1 - ~ i ! ; . ~ o r t c d

bactcria. ijoctcrial nitrificoLion;dcnitr

ificntion. b!icrrihin! 1) rncdiaiivi

i>?ridatinn of mrtnIs.

t:ptakc and

rnctabol ism of

oruanics by plants.

Root cscrctions mny

b c tosic t o

or~nnisms of cntcric origin.

!'ndcr p r n p r r

condi L ions . signi f içant

quantiLics ol t h c s c

contaminants wili bc

:akcn up by plants.

a P = primary effect; S = secondary effect: 1 = incidental effect (effect occurring incidentai to removai of another contaminant)

b metabolism inciudes both biosynthesis and cata bolic reactions

Wetiands Legend

1 Natural Wetlands 2 Constructed Wetiands 3 Shipyard Reference Wetland 4 Gravel Pit Reference Wetland 5 AOSTRA Road Reference Wtland 6 Highway 63 Reference Welland

Figure 1.1 Study sites around the Suncor mine site.

8

Figure 1.2 Aerial photograph of Suncor Lease 86.

1.4 Literature Cited

.Ai berta Environmental Protection ( AEP). 1996. Naphthcnic acids background information: Discussion Report. Environ men tal Criteria Branch, Environmental Assessment Branch. Environmental Reg ulatory Services, Alberta Environmentai Protection. Draft Report . February 29, 1996. 79 pp.

Bishay, F.S. and P.G. Mx. 1996. Constructed wetlands for t h e treatment of oil sands wastewater. Technical Report #5. Prepared for Suncor Inc.. Oil Sands Group by EC'S Consultants Ltd. Sorth Vancouver, British Columbia. 12 Chapters (236 pp.) + Appendices.

Dubuc, Y., P. Janneteau, R. Labonte, C. R J ~ and F. Briere. 1986. Domestic wastewater treatment by peatlands in a northern climate: a w a t e r quality study. Mater Resources Bulletin 22:297-303,

Fan, Ta-P. 1991. Characterization of naphthenic acids in petroleum by fast atom bombardment mass spectrometry. Energy Fuels 2 3 7 1-373.

FTFC (Fine Tailings Fundamentals Consortium). 1993a- Volume 1: Clark hot water extraction fine tailings. In: Advances in Oil Sands TaXngs Research. Alberta Department of Energy, Oil Sands and Research Division. Pu blisher, Edmonton, Alberta. pp. 1-84.

FTFC (Fine Tailings Fundamentals Consortium). 1993b. Volume II: Fine tails and process water reclamation. In: Advances in Uil Sands TaXngs Research. Alberta Department of Energy, Oil Sands and Research Division, Pu blisher, Edmonton. Alberta. pp. 1-52.

Gulley. J.R. and D.J. KIym. 1992, Wetland treatment of oil sands operation wastewaters. In: R., Singhal, A.K. Yehrotra, K. Fytas and J.-L. Collins (eds. ) Environmental Issues and baste .Vanagemen t in Energy and ?linerals Production. Proceedings: Second International Conference on Environmental Issues and Management of kaste in Energy and Mineral Production, Calgary, Alberta, Sep tem b e r 1-4, 1992. A.A. Balkema, Rotterdam, pp. 1431-1438.

Gulley. J.R. and Y. MacKinnon. 1993 Fine tails reclamation utilizing a w e t landscape approach. pp. F23-F24. In: Proceedings of the Fine Tails Symposium. Oil Sands - Our Petroleum F u t u r e Conference. April 4-7. 1993. Fine Tails Fu ndamentals Consortium, Edmonton, Alberta.

Hamilton, S.H. and P.G. Nix. 1991. Wetlands as a treatment system for dyke drainage water: li terature review and feasibility s tudy. Prepared for Suncor Inc., Oil Sands Group by EVS Consultants Ltd., North Vancouver, British Colum bia. 45 pp. + Appendices.

Hamilton. S.H., P.C. Nix and A.B. Sobolewski. 1993. An overview of constructed wetlands as alternatives to conventional waste treatment systems. Fater Pollution Research Journal of Canada 28(3):529-548.

Herman, D.C., P.M. Fedorak, Y.D. HacKinnon and J .k Costerton. 1994. Biodegradation of nap h thenic acids by microbial populations ind igenous to oil s ands tailings. Canadian Journal of .Yic.robiofog~,t- 40:467-477.

Litchfield, D.K. and D.D. Schatz. 1989. Constructed wetlands for wastewater treatment a t A m o c o Oil Company's Yandan, Sorth Dakota ref inery. In: Hammer, D.A. (ed . ) Constructed wetlands for waste water tréa tmen t: municipal. industrial. agricult urai. Proceed ings of the f i r s t international conference on constructed wetlands for wastewater treatment, Chattanooga, TN. Lewis Pu blishers: Chelsea. pp. 233-238.

YacKinnon, M.D. and H. Boerger. 1986. Description of two treatrnents for detoxifying oil sands tailings pond water. hater Pollution Research Journal of Canada 21:496-512.

florea. S., R. Olsen and T. Mildeman. 1990. Passive treatment technology cleans u p Colorado mining waste, bat. Env. Te&. 2:6-9.

Xk, P X - and R.K. Martin. 1992. Detoxification and reclamation of Suncor 's oil sand tailings ponds. En vironmental Toximfogy and Chemistry- 7(2): 171 - 188.

Six, P.C., J.R.P. Stecko and S.H. Hamilton. 1993. Constructed wetlands as a treatment systern for Storm water contaminated by diesel fuel hydrocarbons. Prepared for Energy, Llines and Resources. Canada Centre for ciineral and Energy Technology, Ottawa, Ontario and Imperia1 0i1 Ltd., Burnaby, BC by E l ' s Consultants. Sorth Vancouver, BC. 1 4 3 pp. + Appendices.

Reuter. J.E.. T. Djohan and C R . Goldman. 1992. The use of wetlands for nutr ient removal frorn surface runoff in a cold clirnate region of California - resul ts from a newly constructed wetland at Lake Tahoe. Journal of En vironmen ta1 :Vanagemen t 3 6 S - 53.

Thurston, R-V., R.C. Russo and G.R. Vinogradov. 1981. Ammonia toxicity to fishes: effect of pH on t h e toxicity of the un-ionized arnmonia species. Environmentd Science and Technoiogj- 15(7):837-840.

L'SEPA. 1985. Xmbient water quality cr i ter ia for ammonia - 1984. Criteria and s t anda rds division. CS Environmental Protection Agency. Washington, DC. EPA-440/5-85-001,

Katson, J.T., S.C. Reed, R.H. Kadlec, R.L. Knight and AIE. Ct'hitehouse. 1989. Performance expectations and loading rates for constructed wetlands. In: Hammer, D.A. ( e d . ) Constructed wetlands for wastewater treatmem municipal, industrial. agricultural. Proceedings of the Firs t International Conference on Constructed wetlands for wastewater treatment. Chattanooga. TN. Chelsea: Lewis Pu blishers. pp. 319-351.

2.0 AMMONIA REMOVAL IN CONSTRUCTED WETLANDS RECEIWNG OIL SANDS WASTEWATER

2.1 Introduction

Significant sources of ene rgy are cur ren t ly obtained from mining of oil- rich sands in nor thern Alberta (cur ren t ly more than 15 % of Canada's annual oil needs). Active mining leases dur ing th i s s t u d y were about 26,600 ha, which produced about 300,000 barrels of oil per day (bpd) . Recent approvais f o r new mines include development of another 18,400 ha. Should the cur ren t ly proposed mines become active a total of about 54,000 ha would be dis turbed. These sur face mining p r e s s e s d i s tu rb la rge areas of t h e landscape which will requi re reclamation. Additionally, the extraction process produces large volumes of waste tailings t ha t contain various consti tuents of environmental concern. This s t u d y provides a pilot experiment using constructed wetlands to mitigate the environmental impacts of oil s a n d s mining.

Suncor Inc., Oil Sands Group (Suncor) i s mining oil s ands in nor theastern Alberta, Canada, 25 km north of Fort YcYurray. The mining and subsequent conversion of oil s a n d s bitumen to synthet ic crude oil began in 1967. The Clark Hot Water Extraction Process used to extract bitumen from the oil sands produces a waste tailings. These tailings, made u p of sand, clay, water and unrecovered bitumen. a r e s tored in large tailings ponds. One component of tailings, t h e fine clays known as fine tailings, sett le (i.e., reiease water) ve ry slowly. Therefore, t he re is potential for the reiease of contaminated water as leachate and/or surface r u n-off to terrestrial and/or aq uatic systems as these tailings ponds a r e reclaimed.

A treatment system i s required to mitigate the environmentai impact of leachate and/or run-off water, Both organic (e-g., hydrocarbons) and inorganic (e.g., ammonia) contaminânts are of concern. A constructed wetlands treatment system, which w a s considered possible because of i ts potential for iong-term sustainabiIity and low cost w a s studied. In this s tudy, leachate frorn sa tura ted tailings pond sand dykes (hereaf ter referred to as Dyke water) and tailings pond top water (hereaf te r re fe r red to a s Pond water) were treated in a pilot field-scale mns t ruc t ed wetlands facility .

A varie ty of physical, chemical and biological processes within wetlands are responsible for removing mntaminants. However, there are few examp les of wetlands being used specifically t o treat wastewaters from petrochernical facilities o r o ther sources of organic chemical wastes (Hamilton et al.. 1993). The primary contaminants in t h e wastewaters s tudied were naphthenic acids and amrnonia. However, t h e focus of this chapter was amrnonia removal.

Ammonia i s of concern because of i t s role in eutrophication, increased oxygen demand in t h e receiving environment and i ts toxicity to many aquat ic organisms. Excessive nitrogen loading rnay Iead to eutrophication of t h e receiving environment, especially if' phosphorous loading is already elevated ( Welel, 1983). As ammonia i s oxidized via nitrification approximately 4.6 g of oxygen is consumed pe r gram of ammonium nitrogen oxidized, which cm deplete oxygen in t h e aquat ic environment (Reddy and Patrick, 1984). Ammonia is toxic t o a variety of aquat ic organisms. ünionized amrnonia (NH+ is t h e toxic fraction of total ammonia (NH3 and N H ~ ' ) . The fraction of unionized ammonia var ies with pH and ternperature as does i t s toxicity. Canadian (CCLIE, 1987) and L S (LSEPA, 1985) water q u d i t y guidelines/criteria reflect this. Total ammonia concentration should not exceed 0.93 mg L-' at a pH of 8 and temperature of 20°C fo r t h e protection of aquatic life (CCYE, 1987). Ammonia occurs in oil sands wastewater a t consistently higher concentrations (Le.. 10 - 13 mg L- ' ) than recommended by the water quality guidelines and therefore a treatrnent system is required.

.4 variety of physicat, chemical and biologicd p r e s s e s in wetlands act to translocate and transform nitrogen. Ln wetlands receiving an elevated nitrogen load these processes may act together to effectively rernove nitrogen from the water column (k'atson et al. 1989). Physical, chernicai and biological translocation processes include: particulate set tlin g , diffusion of dissoived forms. sorption of soluble forms on subs t ra te , ammonia volatilization and plant up take. These processes also have opposite/countervailing processes which r e t u r n nitrogen into the water column (e-g. , resuspension, Iitter fall, diffusion). Transformation processes may also remove nitrogen (e.g., nitrXication/denitrification, assimilation) and others may re turn nitrogen (e.g., ammonification/mi~eraiization. nitrogen fixation) to the water column.

It is important to unders tand t h e transiocation and transformation processes acting to remove nitrogen when designing treatment wetland

systerns (Kadlec and Knight, 1996). That is. are t h e con taminants simply retained in the wetlands (e-g,. sequestered in sediments o r plants) and potentially available to re turn to t h e water column. thereby red ucing wetland treatment efficiency. Alternatively. are the contaminants permanently removed from the systern via volatilization and nitrif ication/denitrification. The primary nu trient removal mechanisms are microbial transformation and sedimentation translocation processes. Denitrification is believed to be the primary removal mechanism in wetlands, especially those receiving nitrogen rich wastewater ( H a m m e r and Knight. 1994). Cnder specific conditions plant uptake may be a significant removai

mechanism (e.g., Rogers et a . 1991). while ammonia removal via volatilization may also be significant at pH 8.5 and above (Reddy and Patrick, 1984).

-4mmonia can follow several pathways in wetlands: 1) up t ake by biota. 2 ) immobilization by ion exchange, 3) nitrification by bacteria. Plants can take u p ammonia through their mot systems, as can anaerobic microbes (Mitsch and Gosseiink, 1993). Both convert ammonia to organic matter. Ammonia can be imrnobilized in the sediments via ion exchange onto negatively charged particles (Mitsch and Gosselin k , 1993). Sitrification, the microbial conversion of ammonia to nitrate, is the f i r s t s t ep in the removal

of the ammonia from wetlands. Nitrification occurs in a thin oxidized layer a t the surface of the sediments (Yitsch and Gosselink. 1993) and in the oxidized plant rhizosphere (Reddy and Graetz. 1988). Denitrification. the microbial conversion of nitrate to dinitrogen (X,) (and in lesser amounts nitrous oxide, X 2 0 ) under anaerobic conditions, completes the process of ammonia removal.

Although amrnonia removal was expected to occur in constructed wetlands receiving oil sands wastewater, contaminant fate and treatment efficiency were unknow n. Because nitrification/denitrification are microbial processes

which occur optimally at higher temperatures than usually observed in northern Alberta (57" N). actual removal rates may be slower than observed elsewhere. Nitrification rnay be inhibited by contaminants such a s phenol (Bitton. 1994; Richardson, 1985) which are present in oil sands wastewater. Nitrification may also be inhibited by naphthenic acids above threshold concentrations (Sobolewski and MacKinnon, 1996). In addition. nitrifiers compete with heterotrophic bacteria for various resources such as oxygen (Barnes and Bliss, 1983): and therefore. nitrification may be limited due to cornpetition from bacteria rnineralizing t h e hydrocarbon contaminants. Thus,

in this s tudy, nitrification (hence ammonia removal) may be impacted by t he nor thern latitude, hydrûcarbon contaminants and cornpetition with heterotrophic bacteria.

The objectives of this s t u d y were to determine:

ammonia removal rates under dif ferent wastewater loading rates: - fate of the ammonia removed: and

treatmen t eff iciency of ammonia removal compared w i t h l i terat u re values.

The generaI approach involved monitoring nine constructed wetland t renches over a three year period. Ammonia rernoval was estimated by rneasuring the difference in input and ou tpu t arnmonia concentration and water flow. Contaminant fate was assessed by rneasuring arnmonia input/output Ioads. changes in storage (eg., sequester ing in sediments or plant uptake) and removal processes ( e.g., deni t r if ication and ammonia volatilization). The relative treatment efficiency w a s assessed by comparing observed removal with predicted removal using th ree models.

2.2.1 Site Description and Experimenw Design

Ex perimen fal Fa cilif ,y

The Suncor experimentai constructed wetlands are located in northeastern Alberta. Canada. 25 km north of Fort LicMurray (57"04'S 111°28'W. They are located southwest of Tailings Pond 2 between t h e tailings pond and Highway 63 (F igure 1.1). They were constructed in May and J u n e 1991 and cover an area of approximately 1.4 ha (F igure 1.3). Sine t renches were constructed in 1991 with t h e following characterist ics:

ce11 dimensions - 50 m by 10 m at t h e top and 2 m at the bottom

s ide slope - 2:l t rench slope - 0.5 % (equivalent to a drop of 0.25 m over

50 m) mean depth - 0.35 m

area - 173 m 2

- water volume

liner subs t r a t e

vegetation

- establishment

3 - 52.5 rn (based on 50 m x 3.5 rn x 0.35 m x 0.9 porosity) - 40 mil high density polyethylene - 0.45 m of sand topped ~ i t h 0.13 m of organic muskeg soi1 - 300 Typha lathlia (cattaii) shoots and 60 Scirpus validus (bu l rush) culms (bearing 6 - 15 shoots) were planted in each t rench in J u n e 1991 - 1 year

Experimen Cal Design

Treatrnent waters were applied to the nine t renches in a randomized block

design. Three replicate Control t renches were compared with treatment t renches (Dyke o r Pond). The experimental design changed (Le., flow rates) from year to y e a r (Table 2.1).

.ilmitorhg Programme

There were two sampling schemes depending on the type of anaiysis: 1)

four sampling s ta t ions from inflows to outflows (Le.. A. B. C. D ) were followed for water quali ty and water chemistry, and 2 ) three stations from

inflows to outfiow (Le.. in, mid, ou t ) were followed for sediment and aquat ic emergent macrophyte (hereaf ter re fe r red to as macrophyte) tissue chemistry. In t h e f o u r station sampling scheme, Station A samples were taken from the inflow pipes, Station B w a s located 15 - II metres downstrearn of the inflow to each t rench, Station C 30 - 34 metres downstream of the inflow. and Station D samples were taken from the outflow pipe/weir (50 m downstream of t h e inflow). In the t h ree station sampling scheme, "in" samples were collected from within t h e f i r s t th i rd ( O - 15 rn) of each t rench , t h e "mid" from t h e second third (16 - 30 rn) and t h e "out" from t h e las t third (31 - 45 m).

Sample frequency was limited in 1992 due to water limitations (Le., water was trucked in 1992 rather than pumped as in 1993 and 1994). In 1993 and 1994, water sarnples were cokcted as frequently as weekly to ensure that t h e sample f requency was at least as often as trench turnover (Le., within the estirnated mean hydraulic retention time of 7 to 18 days).

Sedirnent and t issue sample f requency was one to t h ree tintes p e r s t u d y year. In 1994, in situ measurements for ammonia volatilization and denitrification were made two and t h r ee times, respectively.

2.2.2 Sam ple Coiiection

ka ter and Sedimen t

Kater sarnples were collected and shipped in coolers with ice packs to the laboratory the same day to e n s u r e sample analysis was initiated within 24 h. Sedirnent sarnples were collected using a plastic corer (5 c m in diameter) with only the top 1 c m being collected. Several g rab sarnples were collected at each station and placed in "whirl pack bagsn to e n s u r e a minimum of 50 g of sample. Samples w e r e collected and shipped in coolers with ice packs to the laboratory.

Plants

Tissue sam pies were collected f rom Tvpha la tifofia (cattail) and Scirpus vatidus (bu l rush) shoots and roots. A total of t h ree whole plants (Le.. shoot and root) for each species were collected from each trench (i.e., one sample each from each sampling location along t h e t rench) . Shoots and roots were separated, and therefore a total of 12 samples (i.e., t h r ee zones sampled for root and s h o o t for each species) per t rench were collected. Samples were dried at ' 40°C and ground using a coffee gr inder a f te r collection before analys is.

2.2.3.1 Water Chemistry

Three laboratories, t he Lirnnology Laboratory at t h e Cniversity of Al berta, Edmonton, AR; Suncor Laboratory, For t PicMurray, AB and ASL (Analytical Services Laboratory), Vancouver, BC conducted t h e analyses. Most of t he analyses were conducted by the University of Al berta: however. initially t h e other two labs were used. The relative percent difference (RPD) for ammonia within t h e th ree labs was I I %, 3 % and 11 % for t he University of Alberta, Suncor and ASL (Appendix B). The RPD w a s 21 % between ASL and Suncor and University of Alberta and Suncor (Appendix BI. University of Aiberta methods are detailed below, while t h e other iabs used

APHA ( 1989) protocols for the various water chemistry parameters w h ich are detailed in Appendix B.

Arnmonia w a s determined from an unfiltered sarnple us ing the p henoihypochlorite method of Solorzano ( 1969) with the prepared s a m ples being andyzed on the Technicon AutoAnalyzer. Sitrite-nitrate was

determined from a fiitered sampie (0.45 prn HAWP Millipore filter) using the cadmium-copper reduction method of Stainton et ai. (1977). TKS was determined using t h e sulphuric acid-cmpper sulphate rnethod of D'Eiia e t al. ( 1977) and digested samples w e r e analyzed using the Technicon AutoAnalyzer.

2,2,3.2 Sediment and Tissue Chemis try

Exchangeable ammonia and exchangeable nitrite-nitrate nitrogen in sediments were deterrnined using t h e method of Brernner (1965). Total nitrogen (TN) in sediment and t i ssue samples were analyzed using a Control Equipment Corporation Elemental Anaiyzer (240 - XA) follow ing the particulate carbon and nitrogen method of Stainton et al. (19'17).

2.2.4 fn situ Denitrif ication

Although there is no one universall y satisfactory method for rneasuring denitrification (Arah e t al., 1991 ), t h e acetylene inhibition technique was used in this s t u d y (in addition to t h e N mass baiance approach) as it is a relatively easy and rapid technique and has b e n one of t he rnost frequentiy used methods (Seitzinger e t al., 1993). In this study. denitrification was measured in situ modifying methods used in S t r u w e and Kjdler ( 1990) and Chan and Knowles ( 1979).

Field incubations were made in 20 L plastic buckets piaced ups ide down over the sediment surface. The buckets had two sample ports: on top (Le., in t h e bottom of the bucket) and beiow the air-water interface (Le., on the s i d e of the bucket) . Acetylene was supplied at a rate of 340 mL/min for 6 min through 6.35 mm (1/4") teflon hose and dispersed through the water coiumn via a plastic aerator inserted through t h e top sample port. This was the equivalent to supplying approxirnately 2 L acetylene to reach the inhibiting (N20 to N I ) concentration of 10 % b y volume (Chan and Knowles. 1979). Chan and Knowles (1979) found tha t mixing for 5 min was sufficient to disperse t h e acetylene evenly. so dur ing acetylene flow t he hose was

turned by hand near t h e sediment-water interface to enhance mixing. The sample ports were 12.7 m m (1/2") holes plugged with Vaccutainer tube r u bber stoppers. The top s toppers were removed during bucket placement and acetylene addition. Once acetylene addition w a s completed t he hose was removed and t h e port stoppered.

A i r samples were withdrawn following a 24 h incubation period using a 30 mL disposable s y r i n g e and needle through t h e stopper. Samples were then expelled into 22 mL glass vaccutainer tubes with positive p re s su re (Le., 30 mL air samples). Prior to u s e in t h e field, vaccutainer t ubes were opened and resealed following the addition of silica gel and reevacuated (i.e.. re turned to vacuum). Water samples were removed sirnilarly, except a 60 mL syr inge w a s used. M t e r 30 mL of water was drawn into t h e syr inge, 30 mL of helium was ais0 drawn into t he syr inge. The sy r inge was sealed by stoppering t h e needle. Samples were then shaken for 1 minute ( t o equilibrate t h e gases: McAuliffe, 1971) and t h e g a s sample expelled into 22 mL glass vaccuatiner t ubes with positive pressure . Vaccutainer tubes were then resealed with silicone and t ransported to the University of Alberta for analysis by Gas Chromatography within t w o month of collection. Helium blanks were analyzed to estimate background ni t rous oxide in t he gas equilibrated samples ( i.e., water samples).

The acetylene inhibition technique inhibits dinitrogen ( S 2 ) formation and therefore prevents t h e final conversion of n i t ra te nitrogen to dinitrogen (Le., denitrification). The reaction is s topped a t nitrous oxide (\,O) which can be directly measured using gas chromatography techniqut . The S,O

in gas samples w a s rneasured using gas chrornatography with a '%i electron capacity detector (ECD). In addition, car bon dioxide and methane were a k o measured using gas chromatography with FID- Peak areas were determined. Areas were calibrated with known gas concentrations t o estirnate sam pie concentrations based on peak areas.

Amrnonia volatilization w a s predicted using ammonium nitrogen concentration and apparent ammonia volatilization r a t e constant (Freney et al., 1985) for pH 8, 20°C and 5 m / s wind velocity conditions (Kadkc and Knight, 1996). Therefore volatilization may have b e n overestimated because pH was at o r below pH 8, e x œ p t in 1992, mean temperatures were less than 20°C and wind speed estimated from monthly wind r u n s was < l m / s . In addition in

1994, direct measurements w e r e atternpted using phosphoric acid traps- Overturned 10 L plastic buckets were used to isolate a portion of t h e water surface. A 60 mL beaker was filled with 30 mL of 2 % phosphoric acid (Denmead e t al., 1976) as a t r ap for ammonia and placed o n a styrofoam float (1.5 x 10 x 15 c m ) under t h e overturned bucket. The bucket lip was submerged 2 - 4 c m below the water surface to maintain a seâl. Traps were left for several days to accumulate ammonia in t h e phosphoric acid traps. The acid and collected ammonia were analyzed for ammonia as per water sample analysis.

2.2.6 Nitrogen Budget

Nitrogen input and output loads w e r e estimated from input and output nitrogen concentrations and input and output water flows (Appendix C ) . The difference between input and ou tpu t load was assumed to be the load removed o r retained. The load retained (i.e., change in storage) was estimated by measuring nitrogen in several compartments (e-g. , sediments) during 1993 and 1994. The observed nitrogen in t h e sedimen t s and plants in the Dyke and Pond trenches was corrected b y subtracting t h e total in t h e Control trenches to account for background nitrogen and calculate nitrogen storage. The load removed was estimated by measuring the nitrogen elimination processes of volatikation and deni trif ication in 19%. Denitrification in the constructed wetiands trenches was measured and used to estimate loss of nitrite-nitrate nitrogen (Le., loss of nitrogen to t h e atmosphere) after nitrification of ammonia. Amrnonia volatilization from t h e

constructed wetlands was measured and used to estimate loss of ammonia to the atmosphere.

Yass balance equations were determined by estimating t h e various storage compartments together with loss processes. The following mass balance equation was used to estimate t h e removal efficiency of the wetlands under various loading rates:

In years where either compartments or processes were not measured, this equatioo was rewritten to estimate t h e unknown compartment/process.

2.2.7.1 Water Chemistry

The water chemistry data were not statistically anaiyzed beca~ise ot' a Iack of variation (as many control samples and some outflow sarnples had values less than t he detection limit). Data were graphically interpreted to assess whether there w a s a difference between treatrnent waters o r frorn infloh to outflow for a given wastewater. I t was generally obvious when some treatment had o r had not occurred and when t h e treatment observed had o r had not approached conditions observed in t h e Control trenches.

2.2.7.2 Sediment and Tissue Chemistry

Sediment and t issue concentration (mg kg-') data were tested for significan t difference between treatment waters. Ai1 statistical analyses were performed using the SYSTAT statistical package (LI ilkinson. 1990). Foilowing convention, a probability of a 5 0.05 w a s used to define s tatistical sig nif icance,

Repeated measures analysis of variance (.ILVOC'A) tests were conducted o n the total nitrogen and exchangeable ammonia in sediments and the total nitrogen and phosphorous in tissues to a s ses s the effects of *astewater type. Individual t renches were used as replicates of t h e treatment water type ( Le,. the treatments were sampled in triplicate). Yeasurements were replicated three times in each sampling period (Le., date) o r zone (i.e.. in. mid, out) for rnost sediment data. while t issue data kere only replicated spatially not ternporally in each s tudy year. Repeated measures design is

the sarne as a split plot design. with t h e treatments applied to the whole plots or trenches and t h e sampling period and zone applied within plots or trenches. Ko t rench effects w e r e assumed. Post-hoc contrasts were used to determine which rneans were significantly different using Bonferroni adjusted alpha Leveis (i.e., critical level of 0.05 divided by the nurnber of contrasts tes ted) . Testing a subse t of cornparisons rather than ail possible pairwise cornparisons (e.g.. Tukey HSD) often gives more power to the anaiysis (Day and Quinn, 1989: Wilkinson. 1990). Repeated measures ANOVAs were conducted for each parameter (e-g. , total nitrogen in bulrush root, exchangeable ammonia in sedirnents. etc.) and year separately ( Le.. 1993 and 1994).

2-2.7.3 Voiatilization and Denitrification

A T-test between t h e Control and Dyke Drainage treatments us ing t h e t r ench means of data over the two sampling periods and th ree zones was

done instead of repeated rneasures AiiiOV.4. Trench means were used for statist icâl analysis due to the hîgh number of missing data (e-g., volatilization samples lost due to muskrats; ail s ta t ions wi thiri each t rench were not measured for denitrification d u e to t h e length of tirne required to set-up and take samples for each s ta t ion) making other types of sb t i s t i ca l analysis difficult.

2.3 Results and Discussion

2-3.1 Water C h e m i s t r y

In al1 study years ammonia concentrations were reduced from inflow to ouff i ow of t h e treatment wetfands (Table 2.2). Controi inflow concentrations were at least a n o r d e r of magnitude lower than in Dyke and Pond inflows. Dyke inflow concentrations were in t h e r ange of 10 to 15 mg L-'. and Pond inflow concentrations were in the r ange of 10 to 12 m g L-:. Contro i inflow concentrations were generally < 0.05 mg L-': except in 1992 when Ditch water (0.2 mg L-') rather thao Loon Lake water was used. On average the inflow concentrations of Dyke and Pond water were l o ~ e r in 1992 compared w i t h 1993 and 1994. Average Dyke oufflow concentrations were in the range of 0.15 to 12 m g L-'. and Pond outflow concentrations were in the r a n g e of 0.25 to 10 mg L'!. Control outflow concentrations were generally < 0.1 m g L-la On average the outflow concentrations in Dyke and Pond

water were lower in 1992 than 1993; while Dyke outflow concentrations in 1994 were intermediate.

Although ammonia removal w a s s u b s t a n t i d ( ranging from 32 - 99 %), an o r d e r of magnitude difference between the Control t rench outflows and e i the r t he Dyke or Pond t rench outflows remained. In 1992 and 1994, removal rates were in t h e r a n g e of 44 % to 99 %: however, in 1993 rernoval rates were much lower (20 - 40 %) presumably because of t he increased amrnonia loading.

have been impacted by a highly fertilized zone on the tailings pond dyke). In 1993 and 1994. nitrate w a s low (Le.. < 1 mg L-') in both Control and Dyke inflow waters.

Nitrate tended to decrease from infIow to outflow in the Control trenches and increase in the Dyke trenches. The decrease in the Control trenches can be attributed to either organism uptake and/or denitrification (i.e,, bacterial conversion to atmospherical Nî o r N,O). The increase in the Dyke trenches can be attributed to nitrification (Le.. bacterial conversion of ammonia to nitrate nitrogen),

Nitrate was similar between the Dyke and Pond infiows and outflows in 1993

and 1994, although nitrate always increased downstream of Station A but did not rernain elevated. T h e station with the highest concentration of nitrate shifted from B to D over the s tudy season especially in 1994

(Figure 2.2). This shut may indicate a "saturation point" (Le., overloading by ammonia and/or other contarninants, or oxygen limitation for nitrification). The shift in the maximum concentration of nitrate corresponds to the shift in the zone of maximum ammonia removal suggesting saturation. Early in the 1994 s tudy season almost al1 of the ammonia was removed between Stations A and B, while late in the season little was removed between .A and B and the rest Kas evenly removed between B and C and C and D (Figure 2.3). Although these same trends were not as obvious in 1993, perhaps d u e to the overloading of the wetland trenches. the proportion of the total ammonia removal generally decreased between Stations A and B.

Total hj'eldahl ~Vitrogen

Total kjeldahl nitrogen ( T U ) is a measure of organic nitrogen plus ammonia-ammonium nitrogen. The ratio of TKiV to ammonia nitrogen in the water of both Dyke and Pond trenches was close to 1 which suggests that very little of t h e TKN w a s organic. There was no change or a slight decrease in this ratio from inflow to outflow suggesting that no more organic nitrogen w a s leaving the wetlands than was entering. TKN in the Dyke and Pond outflows was higher than in the Control outflows refiecting the differences in ammonia nitrogen (Table 2.2). In the Control wetlands, ammonia was only a mal1 percentage of TKN (< 5 %).

Exchangea ble .4mmonia

Exchangeable ammonia concentrations were tower in the sedirnents of t h e Control trenches than ei ther t h e Dyke or Pond trenches (Table 2.4: Figure 2.4). These differences w e r e significantly different in both 1993 ( F = 85.7,

p < 0.001) and 1994 ( F = 25.0. p = 0.008). In addition. t r ends over time and distance from t h e inflow were observed.

Trends of increasing ammonia with continued ammonia loading were

significant in 1994 ( F = 4.9. p = 0.042) for Dyke trenches (F igure 2.4). In t h e Dyke trenches, exchangeable ammonia increased in 1993 from 14.8 mg kg-' in June to 34.4 mg kg-' in September. and in 1994 it increaçed from 9.6 m g kg-' in June to 13.6 mg kg-' in August to 20.1 mg kg'' in Septernber. Arnmonia accumulated over time in the sediments. but when wastewater (ammonia) application w a s discontinued in t h e autumn and w inter. ammonia concentrations in t h e Dyke t rench sediments decreased. For example between 1993 and 1994, the autumn 1993 sedirnent ammonia concentration of 34.4 m g kg-' decreased over winter to 9.6 m g kg-' by June 1994.

Trends of decreasing arnmonia from inflow to outflow were significant in 1993 ( F = 63.7. p < 0.001) and 1994 ( F = 12.3. p = 0.004) for Dyke t renches (Figure 2.5). In the Dyke trenches. exchangeable ammonia decreased in 1993 from a mean of 30.8 m g kg-' at the inflow to 18.6 mg kg'' at t h e outflow. and in 1994 i t decreased from 22.9 mg kg'' a t the inflow to 8.1 mg kg-' at the outflow. The decrease in amrnonia in the sediments downstream of the inflow reflected changes in water chemistry (Le.. ammonia is removed from the water column at upstream locations: therefore. t h e r e is less to en te r the downstream sedirnents). In addition. t he higher concentrations observed in 1993 were likely d u e to the higher ammonia loading.

Exchangea ble %ifrate

Sedinent nitrate was very low compared with ammonia (Table 2.4). This was expected since 1) nitrate inputs were low (Le., often less than detection iimits), and 2) ni t rate formation from ammonia is the rate limiting step in ammonia removai. Nitrification, t he microbiai process mnver t ing ammonia to nitrate. must occur prior to denitrification. Denitrification, the microbial process converting ni t rate to dinitrogen which is released to the

atrnosphere, is usually limited by nitrification (ilitsch and Gosselin k. l993). Therefore. low ievels of nitrate were expected in the sediments, because as nitrates w e r e formed they would diffuse to t h e anaerobic zones where they would b e converted to dinitrogen. Nitrate was higher in t h e Dyke trenches compared w ith the Control trenches (Table 2.4): however. statistical analysis was not conducted due to the large number of values which were less than the detection limit (DL). Nitrate w a s less than the DL in al1 trenches in J u n e and was generally higher in August than Septem ber.

Total .Vitmgen

TX in the sediments was not significantly different among treatments (Table 2.4). No t rends over time o r distance downstream of the inflow were observed. The ammonia observed in the Dyke trenches was only about 0.01 to 0.02 % of the TX, which w a s not a significant proportion of the N in the

sediment s torage pool. In fact. the proportion ammonia w a s indistinguishable when TN in the Dyke trenches w a s mmpared w i t h Control trenches making it difficult to quantify the nitrogen load retained in the

sediments (as measured by Th').

In 1994, the ratio of total carbon to total nitrogen was 26.6:l and 28.6:l for

the Control and Dyke trenches, respectively, while the ratio of total nitrogen (S) to phosphorous (P) was 6.6:l and 5 5 1 for both Control and Dyke trenches. respectively . This suggests that signif icant total S

accumulation in the sediments may not have occurred.

2.3.3 Macrophyte Tissue Chemistry

TL concentration in the macrophyte tissues collected from t h e Dyke and Pond trenches tended to be elevated cornpared with the Control trenches (Table 2.5; Figure 2.6). En 1993, TS was significantly different among the three treatments for both cattail ( Typha latifolia) shoot ( F = 10.4, p = 0.011) and m o t (Le,, rhizome and root tissues: F = 10.9, p = 0.01) tissues. However, TN concentration w a s only different in bulrush (Scirpus validus) root t issues (F = 29.7, p = 0.001) and not bulrush shoot tissues. Mhen differences were significant TN concentrations in t issues from Control t renches were iower than in tissues from the Dyke and Pond trenches. Although TN concentrations were generally higher in Dyke trenches than Control trenches in 1994, TN w a s only significantly different in cattail r m t t issues (cattail roots p = 0.001; bulrush roots p = 0.2).

Root TN concentrations tended to be higher in trenches receiving the higher ammonia loading (i.e.. Dyke and Pond trenches), more so than shoot concentrations. Ratios of N:P in t issues fur ther supported these trends. Shoot N:P ratios for cattails (Control 10:1, Dyke 8:l) and bulrush (Control 9.49 , Dyke 1051 ) were sirnilar between treatment waters as would b e expected since there w a s no statistical difference between S tissue concentrations. However, rmt N:P ratios for cattails (Control 4.4:1, Dyke 7 - 2 1 } and bulrush (Control 5.4:1, Dyke 7.5:1) were higher in the tissues from the Dyke trenches suggesting uptake of S. N was taken up by t h e mots in trenches with higher concentrations of ammonia in the water and sediments (Le., Dyke and Pond trenches). However, when the concentration data were corrected for biomass in each sarnpling zone these differences were no longer significant in 1993 (Figure 2.6) In 1994, these differences may have been significant (Figure 2.6): however. the data were confounded by muskrat "eat out" in the Control trenches.

Potential denitrification ( e . the conversion of nitrate to nitrous oxide/dinitrogen) was measured in situ using the acetylene blockage

9 technique in 1994. Potential denitrification (Le., mg N,O-N evolved per m' per day) was higher in the Dyke trenches than the Control trenches (Figure 2.7). 'leans for the Control and Dyke trenches over the four

n

sampling periods and three zones were 0.04 m g .\:/m'/day and 8.16 mg ~ / r n ' / d a ~ . respectively. These rneans w e r e significantly different ( T =

-8.042; p = 0.015). However, the measured denitrification rate of 0.19 k g N/wetlands/season accounted for only 4 % of the 4.8 k g arnmonia N/wetlands/season removed in the Dyke trenches. This observation may be d u e to the limitations of the technique rather than the lack of

occurrence of denitrification (Seitzinger et al., 1993; Dalsgaard and Bak, 1992).

The acetylene inhibition technique rnay underestirnate denitrification rates. The acetylene inhibition technique fails to cap tu re coupled nitrification/denitrification due to the inhi bition of nitrification by acetylene (Seitzinger et al., 1993). No additional nitrate was added during the incubation period, onIy the nitrate present was available for denitrification. Consequently, denitrification as measured could have been underestirnated. Standing stock of nitrate in the sediment and w a t e r were

2 2 estimated as 1.23 m g / m and 42.9 mg/m based on rnean nitrate nitrogen

concentrations of 0.1 mg kg-' and 143 pg/L. reçpectively. Assuming t h e estimated standing stock of ni t ra te w a s present within t h e chamber du r ing the incubation period, then t h e measured denitrification should have t e e n higher. Acetylene may not reduce N 2 0 in t he presence of some organisms which would lead to a n underestirnation of in situ ra tes (Dalsgaard and Bak, 1992). Additionally, t h e concentration gradient of K,O may change as N20 increases inside t h e cham ber. The theoretical constant gradient i s linear: however. t h e expected gradient will decrease over t i m e (Lem ke. 1997). If linearity i s assumed then a downward bias of 34% may occur (Anthony et al., 1995). Hutchinson and Mosier (1981) developed an adjustrnent for this underestimate; however, t h e assumption required for t h e calculation only holds for a subse t of samples (Anthony et al., 1995). Therefore. linearity was assumed, which could also have led to a n underestirnate of denitrification.

Nitrate loss measured as denitrification in other s tudies ranged frorn 60 t o 70 % in unvegetated rnicrocosms in a treatment wetland (Cooke. 1994). 73 to 90 % in a flow-through Phragmites/gravel s u bsurface wetland (Stengel et al., 1987). to about 90 % in soi1 microcosms in a treatrnent wetland (Bartlett et al., 1979). These values are much higher than those measured in t h i s s tudy (about 4 %: see later) . Assuming the still unaccounted for ammonia was indeed nitrified/denitrified. then the proportion (about 87 %)

would be comparable w ith denitrification measured in other s tud ies (e.g.. Stengel et al., 1987). However. th is may not be valid given the cooler temperatures (e-g., Brodrick et al., 1988).

The rates observed in both Control and Dyke trenches were lower than t he r ange of 60 to 87 mg m-' d-' observed in eight riparian wetlands (Seitzinger 1994). Although t h e riparian wetlands studied were selected to include various degrees of anthropogenic N inputs, N loading was not quantified (Seitzinger, 1994) making comparison with this s tudy difficult. An average denitrification r a t e for marshes about 100 km north of

2 Edmonton were 47.1 mg N/m /day (Mewhort, pers. corn.). These ra tes were lower than those obtained by Seitzinger (1994). which were obtained unde r a higher incubation temperature. These denitrification ra tes probably reflect systems receiving a lower N load than in t h e Dyke trenches.

Seitzinger et al. ( 1993) has determined that t he acetylene inhi bition technique underestimates rates by 50 % due to failure to prevent reduction of N 2 0 and another 35 % due to inhibition of nitrification. However. even

assurning this degree of underestimation. t h e denitrification r a t e w a s still only 17.4 mg m-' d-', mnsiderably lower than 200 mg m" d-l of S rernoved from t h e water. Either another removal pathway was dominant or perhaps under higher N loadings/denitrification rates, the proportion underestimated increases. Aiternatively, proportionately higher numbers of organisms that a r e not inhibited by acetylene were present (e.g., Daisgaard and Bak. 1992) than in Seitzinger e t a1.k (1993) s tudy.

Loss of ammonia via volatilization (as measured by amrnonia trapped in 2

% phosphoric acid solution) was variable. Control trenches exhi bited very Iittle arnmonia volatilization compared w ith Dyke trenches (Figure 2.8). Heans for the Control and Dyke t renches over t h e two sampling periods and th ree zones were 0.025 t 0.039 mg ~/rn'/day and 0.185 I 0.060 mg N/mL/day, respectively. These means were significantly different ( T =

-4.01; p = 0.016). The measured loss of amrnonia via volatifkation was about 0.004 kg trench" season-' in the Dyke trenches or < 1 % of the ammonia removed. Ammonia volatilization as measured, may b e an underestimate. The buckets used to t r ap voiatilizing gases from t h e water surface maÿ also prevent wind action that could have accelerated the voiat ilization rates. kind s p e e d , p H temperature and ammonia concentration are the governing factors affecting ammonia volatilization (Freney et al., 1990).

.A volatilization model was used to predict t he voiatilization rates. because the measured ammonia volatilization rates could have been a n underestimate. The predicted rates w e r e compared with the measured rates to assess the potential for significant ammonia removal via volatilization. Predicted ammonia volatilization (from a model developed by Freney et al.. 1983) w a s one to two orders of magnitude greater than measured volatilization. The mode1 equations were cornbined to calculate a n ammonia emission ra te constant based on t h e ammonia concentration in the water as follows (Kadlec and Knight, 1996):

where: J, = ammonia emission rate. g N m -2

ka, = apparent ammonia volatilization r a t e constant. m/d CU,-+ = unionized ammonia in liquid phase at the water surface.

g/m

The apparent ammonia volatilization r a t e (k,) constant of 2.39 m/y (or 0.0063 m/d) was used. This mode1 assumes a pH of 8, temperature of 20°C and wind speed of 5 m/s . This rate constant would likeIy a overestimate ammonia emission rates because the mean water temperature and pH w e r e lower at 17.7"C and 7.6, respectively. Although t h e wind speed w a s not actually measured, the average w ind speed was calculated f rom monthly wind r u n s to range from 0.4 m/s t o 0.8 m/s, which was lower than the wind speed used (5 m/s) to calculate the K,. K, values were only available for wind speeds of 3 m/s. L'sing t h e K, of 0.0065 m/d and assurning 0.57 mg L" unionized ammonia (i.e.. 3.82 % of 15 mg L-' total ammonia at pH 8 and 20°C is unionized; LSEPA 1979), t he ammonia emission ra te from t h e Dyke trenches was calculated a s 3.7 m g rn-' d-'. Volatilization w a s not a significant pathway, because pH w a s not consistently above 8.5 (Reddy and Patrick. 1984). Diurnal sarnpling was conducted in ear ly September of 1994 and pH increased in t h e late afternoon: however, on average it remained below 8.5.

Civen this and tha t t h e assumptions are overestirnates of average in situ conditions, it i s possible tha t t h e actual r a t e is approximated better by in situ acid t r a p s than by t h e above mode!. Although t h e calculated volatilization rate i s significantly grea te r than t h e measured rate it still would only account for less than 3 % of t h e ammonia removed from t h e D y k e trenches. Neither t h e measured o r predicted ra tes suggest tha t ammonia volatilization played a significan t role in nitrogen removal f rom t h e treatrnent wetlands,

2.3.6 Nitrogen Mass Balance

The mass balance for ammonia nitrogen in t h e constructed wetlands treatment t renches w a s determined for each s t u d y year (Table 2.6)- Ammonia nitrogen in inflow and outflow water, change in s torage and microbial removal processes were measured, and a mass balance for ammonia

nitrogen in t h e Dyke and Pond t renches was determined. In 1994. the only yea r with a mmplete budget, the difference between inflow and outflow (i.e., the amount of ammonia "removed" o r "retained" by t h e wetlands) was 4.8 kg trench-l season-' (206 mg m-' d-I). while t he sum of t he measured removal processes and change in nitrogen s torage w a s only 2.35 k g t r e n d season-' (101 mg m-' d-'1.

Assuming t h e source of aii the ammonia rneasured in t he sediments was from the wastewater and not ammonification. then a significant portion of t h e input load w a s sequestered in the sediments. There was s o much

organic N (as rneasured by TX) in t he sediments in both t h e Control and Dyke t renches that ammonia retention could not b e detected. In fact, TN w a s higher in the Control (1.2 kg ITI-~) t renches than Dyke (0.9 kg m-') trenches. However, arnmonia removal to t h e sediments was about 101 m g m.' d-I (37.4 % of input nitrogen) for a total of 13.384 mg m-' (Figure 2.9). This accumuiation may be an overestirnate s ince the amount of ammonia present a t the beginning of t he s tudy year was assumed to be no more than what was in the Control t renches at t h e end of t h e s t u d y year (cf. Section 2.3.8). Conversely. th is i s only about 1.4% of t h e TS. which is within the e r r o r of t h e TN analysis. and therefore the measurement of T'i in sediments may not accurately ref lect the ammonia nitrogen accumulation during the s tudy . Additional1~-. TN anaiysis may fur ther underestimate ammonia accumulation because incorporation into organic matter would also not have been detected. Removai of ammonia nitrogen to the sediments was at least 49 % (2.3 kg t rench-i season*!: 101 mg m.' d- ' ) of t h e amrnonia nitrogen retained (more if immobilization occurred ). w hile plant uptake accounted for approximately 2 % (0.10 k g t r e n ~ h - ~ season-' o r 4.5 mg m.'

Sitrogen removal b y plant uptake is not considered to be a main rernoval 1 '

processes. Although plant uptake may reach 500 m g rn= d" annually (van Oostrom and Russell, 1994). t h e majority of nutr ients are leached back into the water at the end of each growing season (Richardson and Nichols. 1985). Klopateck (1975. 1978 cited in Richardson and Nichols. 1985) found

'l

t ha t 48 mg N/mA/d was translocated from wetlands soils to plants over t he yea r and that about 42 % of this nitrogen was leached back into t h e water column at t h e end of the growing season. Plant uptake in a treatment systern in California only accounted for 12 - 16 % of the nitrogen removed (Gersberg, 1986).

Volatilization and denitrification accou nted for approximately 0.1 % (0.004 k g trench" season-' or 0.22 mg m-' d-l) and 4 % (0.19 kg trench-' season-' o r 10.6 mg rn-2 d-l) of the amrnonia retained. respectively. Volatilization of ammonia rnay account for loçseç up to 100 mg m.' d-' at pH 8 (based on an ernpirical equation developed for pred icting ammonia volatilization from flooded rice fields by Freney et al.. 1985); however. based on the environmental conditions observed the mode1 predicted that only 3.7 mg m-' d-' of ammonia could have been volatilized. Althoug h denitrif ication is often assumed to b e the primary removal process (Nichols, 1983). in t h i s s tudy it may only represent a t most 50 % of the observed removal. However, denitrification (as measured ) could not be verified experirnentally to be a predominant pathway possibly due to the limitations of t he technique rather than the lack of occurrence of denitrification (Seitzinger et al., 1993; Dalsgaard and Bak, 1992).

Although the removal processes and storage compartments measured did not account for al1 the ammonia nitrogen retained by the wetlands, these processes and cornpartments were expected to be responsibIe for the losses observed because there was no other removal pathway. It is unlikely that removal to the vegetation was significantly u nderestimated. R e m o v a l to the sediments may have b e n underestimated bemuse incorporation of ammonia into organic matter would not have been detected by TS measurements. Potential denitrification rates may also have b e n underestirnated. Although volatilization measured in situ may also be a n underestimate because of reduced wind action over the sample site, it w a s not assumed to be a significant removal pathway because predicted rates were aIso 10%.

In estuarine constructed wetlands the dominant removal pathways include both accumulation in organic matter in the sediments and denitrification. Under lower nutrient loading the ratio of removal to the sediments and via denitrification is about 1.4:1, w h i l e under higher nutrient loading both processes increase though denitrification more so. such that the ratio flips to abou t 1:3 ( C r a f t , 1997). Therefore, sediment retention and nitrification/denitrification were assumed to b e t h e dominant ammonia removal path ways here also.

2.3.7 Treatment Eçf iciency

The capability of constructed wetlands, receiving oil sands wastewater, to remove ammonia was assessed by comparing observed effluent ammonia

concentrations with predicted concentrations using th ree differen t models (Figure 2.10: Appendix D}. Model 1 is based on nitrification rates in attached growth biological reactors (Reed et al., 1994). Yodels 2 and 3 are based on ammonia removal in various wetlands. Model 2 is based on natural wetlands and free surface and submerged bed constructed t~e t l ands (WPCF. 1990) and Yodel 3 is based on 17 free surface wetlands (Harnmer and Knight. 1994). Since none of these models reflect only f ree surface constructed wetlands in northern latitudes, al1 three were exarnined to better understand the relative treatment efficiency of the constructed wetland trenches receiving Dyke and Pond water.

In general. the observed effluent ammonia concentrations from the constructed wetland t renches were significantly different from the predicted concentrations in 12 of the 15 data sets (Figure 2.10). For 8 of the 12 data sets, the observed treatment signif icantly out performed the predicted treatment, whîle predicted treatment outperformed the observed treatment in four data sets.

Yodel 1 assumes that constructed wetlands are attached growth biological reactors and that their performance can be estimated by first-order plug- f l ow kinetics and is based on data for available sur face area. hydraulic retention times and water temperature (Reed et al., 1994). Observed ammonia removal in 1992 - 1994 overlapped with t h e ammonia removal range predicted by Yodel 1. Since Yodel 1 reflects nitrification rates t h e sirnilarity between the observed and predicted data indicated that t h e primary ammonia removal mechanisrn was likely nitrification/denitrif ication. Significant differences between three of t h e five observed and rnodelled data sets were noted: however. t he observed outflow arnmonia values were lower than the mode1 values for two of these data sets (Figure 2.10).

Although these differences may not be biologically significant, they can be explained by experimental error, the effects of mineralization, short-ierm p tant uptake or amrnonia volatilization, The constructed wetland t renches removed at least as much ammonia as expected from Model i in three of t h e five data sets ( Le., significantly higher o r not significantly different ). This sugges ts that the performance of t h e t renches was neither inhibited by the petroleum hydrocarbons nor the northern location.

Model 2 is based on a regression analysis of data €rom constructed and natural wetland treatment systems (WPCF. 1990) and is based on hydraulic loading rate. Observed ammonia removal in 1992 - 1994 overlapped with

t h e ammonia removal range predicted by Model 2. However, significant differences between the observed and modelled data were noted for al1 yea r s and treatrnent waters. The observed outflow ammonia values were generally comparable with model values except in 1993 where observed values exceeded predicted values s u bstantially. These d ifferences w e r e Iikely due to the source data used in the derivation of -iode[ 2. Although t h e input arnmonia concentrations and loads fi t in the range of t h e treatment wetlands used to derive t h e model, the inclusion of da ta from submerged beds may have increased rernoval efficiencies beyond that possible by t h e free surface wetlands used a t Suncor. The cornparison of predicted data from Mode1 2 with t h e observed data sugges ts that ammonia removal at Suncor may be somewhat inhibited by the petroleurn hydrocarbons or the northern location. Temperature was not included in this mode1 and may have had an effect s ince many of the wetlands used to derive Mode1 2 were from t h e southern United States. However, Mahli and McGill (1982) observed tha t optimum nitrification rates change with climate. s o in cooler climates a lower optimum temperature for nitrification OCCUîS.

Yodel 3 is based on a regression analysis of' data from 17 free sur face constructed wetland treatment systerns ( H a m m e r and Knight, 1994) and is based on ammonia loading. Observed arnmonia rernoval in 1992 - 1994 overlapped with the ammonia rernoval predicted by ?Iode1 3. However, significant differences between t h e observed and modelled data were noted for al1 years and treatment waters; however. t h e observed outflow ammonia values were alwaÿs lower than the predicted values. Although these differences may not be biologically significant, they can be explained by experimental e r ror , and differences between systems with respect to mineralization, short-term plant uptake or ammonia volatilization. The constructed wetlands removed as much ammonia as expected from Yodel 3 and sugges ts t h e performance of the t renches were neither inhibited by t h e petroleurn hydrocarbons nor t h e northern location.

Cornparisons of t h e treatment wetlands with the rnodelled data (in t h e three models) sugges ts that the treatment efficiency in the Suncor constructed wetlands was sornewhat lower than systems receiving similar amrnonia loading under the conditions for Models 1 and 2, but comparable and even bet ter in some cases than for Model 3. Model 3 was most representat ive of t h e conditions in this s tudy, and the observed data were comparable or better than resul ts predicted by this model. Therefore, ammonia removal

(Le., primarily nitrification/denitrification p r e s s e s ) w a s not impacted by t h e northern lat i tude of the s tudy site. petroleum hydrocarbon contamination, and/or cornpetition for resources.

Although this was only a th ree year s tudy, there i s potential for long-term treatment. Ammonia removai was consistent arnong t h e th ree years s tudy . with the highest Ioad being removed in 1993, although th i s was not the year w ith t h e hig hes t eff iciency . Linlike phosp horous, wetlands a r e expected to retain t h e rate of nitrogen removal and are likely t o improve with age (Craft. 1997).

2.3.8 Long-term Treatment

The dominant removal mechanisms in this s t u d y were assumed to be nitrification/denitrification a n d s e d i m e n t r e t e n t i o n . Nitrification/denitrification completely removes nitrogen from t h e wetland to t he atmosphere and is sustainable in t h e long-term (Craft. 1997: Johnston, 1991 ). Sediment retention occurs via sorption. and sedirnent and organic mat ter accretion. AIthough sorption i s revers i ble. sediment and organic matter accretion usually are not (Johnston. 1991 ).

The amrnonia in t h e sediments was likely retained by sorption and therefore w a s potentially reversible. Ammonia measured in t h e sediments decreased from the end of t h e 1993 year to t h e beginning of 1994 (Figure 2.4). Lnfortuneatly wastewater application had begun about t h r e e w e e k s prior to t he June sarnple collection in 1994. ffowever, if linear accumulation is

assumed then sedirnent concentration increased about 0.106 mg kg-' per day, and the initial concentration waç estimated to be about 6 mg kgm'. This initial concentration was considerably lower than the 34.4 mg kg-' a t t h e end of 1993.

A similar reduction was observed in t h e Natural Wetland and Zeference Wetlands between autumn and spring/surnmer. Ammonia nitrogen in t h e Katural Wetland sedirnents decreased from 41.9 m g kg-' in Septem ber 1993 t o 15.6 mg kg" in June 1994. while in the Reference Wetlands it decreased from 5.3 mg kgml to 3.0 mg kg-', respectively. In Septernber 1994. ammonia nitrogen increased t o 20 and 9 mg kg-' in t h e Natural and Reference Wetlands, respectively. The decrease muid be due to microbial processes, and dilution o r f lushing from t h e spr ing runoff. Regardless, sediment

retention capacity was renewed, but ammonia release may have occurred in the spring prior to the start of t h e experiment.

Assurning t h e unaccounted for amrnonia removal was nitrifiedjdenitrified. then the constructed wetlands should be capable of removing a t least 39% of ammonia inputs indefinitely. .Mternatively. it could have been incorporated into microbial biomass and not distinguished by measurements of TN in the sediments. Further removal would be dependent on the retention capacity of the wetiand sediments, If the sorbed ammonia could be incorporated into microbial biomass and immobilized in the sedimen ts then sediment retention could also be indefinite. Denitrification and organic matter accretion Vary with wetland age. A s the wetiand ages in the f i r s t decade or so these processes tend to increase and then stabilize (Craft, 1997), suggesting ammonia removal could occur at the observed rates indefinitely under similar ammonia loading rates.

2.4 Conclusions

Average ammonia removal rates (Le., mass reductions) ranged from 42 % to

99 % for Dyke water and 32 % to 96 % for Pond water during the three year study. These rates were comparable with rates reported in the l i terature (Nichols, 1983, Knight et al.. 1985. Katson et al. 1989). These removal rates varied depending on ammonia nitrogen loading ra tes (Figure 2.1). Year 1 (1992) had a relatively low loading rate that resulted in a high treatment efficiency (Le., percent removal). Year 2 (1993) had a very high loading ra te that resulted in a f ow treatment efficiency. Year 3 ( 1994)

had an intermediate loading rate that resulted in a n intermediate treatment ef f iciency .

Ammonia loading rates ranged from 1.8 to 7.9 kg/ha/d. Assuming 1994 water chemistry conditions, loading rates should not exceed 3 k g N/ha/d to maintain effective rernoval (i.e., about 70 % mass rernoval and a n outflow concentration of < 2 mg N/L). Loading rates and outflow concentrations observed in other studies ranged from 0.1 to 26.4 kg N/ha/d and 0.2 mg L" to 48 mg L-l, respectively (Harnrner and Knight. 1994). Hamrner and Knight (1994) suggest that loading rates should not exceed 3 - 5 kg S/ha/d to produce outflow concentrations of < 4 mg S/L. Removal ra tes for Pond water tended to be lower than those for Dyke water: however, similar rates were not expected due to the higher contaminant load (e.g.. hydrocar bons, suspended solids) in t h e Pond water.

The dominant removal path ways w e r e suspected to be sed irnent re ten tion and nitrification/denitrif ication. In 1994. al1 removal path ways w e r e measured for Dyke water. Ammonia removal to the sediment and t issue

storage pools accounted for about 49 % and 2 % of total rernoval. res pectiveiy . Ammonia removal via volatilization (as measured ) accounted for less than 1 % of total removai. Although volatilization may have been underestimated d u e to the field methodology, it is not likeiy that amrnonia removal via volatilization exceeded 3 % of t h e total ammonia removed based on predicted rates. Although t h e measured in sif u denitrification rates only accounted for 4 % of t h e total ammonia removed, i t is likely that nitrification/denitrif ication was a dominant removal mechanism.

.4lthough ammonia removal w a s dependent on input load, t he observed ammonia concentrations in t h e outflow water were within the range predicted in several models. Ammonia was effectively removed by constructed wetlands receiving oil s ands wastewater when compared with o ther free-surface wetlands. despite t h e other consti tuents in the wastewater ( i.e.. hydrocarbons) and being situated at a northern latitude. Observed data were comparable w i t h or better than predicted data from the mode1 representing amrnonia rernoval in Pree-surface wetlands.

Table 2.1 Experimental design for the Constructed Wetlands study. 1992- 1994.

- - - - - - - - - -

Trench Treatment Design Hydraulic Duration Nu m ber Cvater Flow Rate Loading

(L /min ) Ra te (cm/d) Start Stop

Study Year 1992

controia

Dyke

Pond

controib

Dyke

Pond

Pond

controlb

Dyke

D y k e

D y k e

D y k e

Dyke

D y k e

Dyke

Study Year 1993

June 29

June 29

June 29

June 19

June 4

June 19

June 19 Aug. 9 Sept. 1

Yay 18

Yay 18

J u n e 8 AU^. 15-18

June 8 Aug. 15-L8

J u n e 2

June 2

June 2

June 2

Sept. 27

Sept. 27

Sept. 27

Sept. 28

Sept. 28

Sept. 28

Aug. 9 Sept. I

Sept. 28

Sept. 28

Sept. 28

.Aug. 13 Sept 28

Aug. 13 Sept. 28

Sept. 28

Sept. 28

Sept. 28

Sept. 28

n In 1!392watcr rram a ditch parallcl ta t h e study s i tc was uscd. which k a s n mixLurc o f suri'acc runoff. ground water and possibly dykc drainage uatcr i i . c . . Dykr watcr).

b In 1993 and 1994. Lhc d i t c h was not uscd becausc of thc possiblc influcncc of Dykc watcr . Thc source was changed to thc sou th minc drainagc pond ( i . c. . Loon Lakc ) which mccts watcr quality guidc!incs to bc rclcased to Lhc .\thabasca Rivt lr .

c Split trenchcs wcrc combincd ( i - c . . inputs and outputs wcrc sununcd! for cach trcnch for discussion os dividcr intcgrity was questionablc. Scdimcnt and Lissuc chemistry. denitrirication and voiatiliza~ion werc not measured in the Split trcnchcs.

(1 Trcnch 3k and 3E reccivcd phosphate additions. Trcnch 3k kas also a c r a t c d .

Table 2.2 Ammonia- . nitrite-nitrate- and total k jeldahl n i trogen concentrations (mean) for the i n f l o w water (-4) and outflow water (D) of t he mnstructed wetlands t renches.

k'ear Treat ment Station Nitrogen (mg L-') kater

Amrnonia Nitrite- Total nitrate KieIdahl

1992 Control A 0.21 0.96 1.7

D 0.09 0.03 1.2

Dyke -4

D

Pond A

D

1993 Control A

D

D ~ k e .A

D

Pond A 13.3 0.0'7 18.1

D 10.4 O. 4 4 L4.9

1994 Control -4 0.00 1 0.002 O. 4

D 0.00 1 0.0004 0.6

Dyke -4 14-6 0.008 16.2

D 3.7 O. 1 3.9

Table 2.3 Ammonia removal in constructed wetlands receiving oit sands wastewater.

Year/ Ammonia Inputs Ammonia Outputs Removal Load Water (mean 2 S D ) (rnean 2 S D ) ( % of Hemotred

1992 Dy k e 1 1.622.8 185 0.1750.17 2 99 183 Pond 10.753.7 187 0.4410.4 1 7 96 180

1993 Dyke 16.253.3 787 11.8+2.0 453 42 332 Pond 13.223.8 607 10.952.9 4 12 32 193

1994 D y k e 14.653.0 271 3.722.5 65 76 206

Table 2.4 Sutrient sediment chemistry (mg kg- j ) in the constructed wetlands trenches, Septcm bcr.

Year Treatment Exchangeable TC TN TF TS:TP tjater Nitrogen

-

SH,' %O?*

1993 Controi 3.4 nma 43000 1320 nrn nrn

Dy ke 34.4 nm 38000 1220 nm nm

Pond 38.8 nm 36200 1090 nrn nm

1994 Control 2.11 0.06 40300 1690 249 6.6

Dyke 20.1 0.10 36300 1275 230 - - 3.3

a nrn = not measured

Table 2.5 Concentration of nutrients in plant tissue chemistry (mean) in t h e constructed wetlands trenches.

-- -

Species Year Treatmen t Tissue Total Total hater Type Sitrogen P h o s p h ~ r o u s

W (mg kg-' dry

Cattail 1993~ Control shoot 0.6'7 n rnC

root 1.01 nm

D y k e shoot 0.77* nrn

Pond shoot

root

1994~ Control shoot

root

Dyke shoot

root

B u l r u s h 1993 Control shoot

root

nm

nm

1308

lS'8

1154

1478

nm

nrn

D y kt3 shoo t 0.97 nrn

rmt 2.221: nm

Pond shoot 0.83 nm

root 2.27* nm

1994 Control shoot 1.42 1308

root 0.93 1726

D y k e shoot 1.45 1537

root 1.90 2306

a September 21, 1993 b August 17, 1994 c nrn = not measured ~t asterisk indicates statistically different from the Control (a =

0.05) for total nitrogen

Table 2.6 Ammonia nitrogen m a s balance ( k g t renck i season-') for D y k e and Pond water.

1992 1993 1994 Parameter r

~ y k c ~ ponda Dykc ~ontl ' D'~P" I

Input fioading) k g amrnonia S t r e n d season-' 2.9 3.0 16.0 12.3 6.3

a af'ter 90 days of wastewater application: ! kg t r e n ~ h - ~ rnson'' = 63 a{ a*' -7 d - ! - 1 ' af ter 116 days of wastewater application: 1 kg trench" s e a s o d = 19 ig a ; d A ' after LOI days of wastewater application: i kg irench-f rearon-' : ii q a-f d - f after 133 days of wastewater application: i kg lrench-' seasofi = 13 i g 6' d-

e a s measured in situ The sum of al1 rernoval to storage pool (çediments and uptake b y p l a n t s ) a n d b y p r o c e s s e s ( t -o lat i l i zat ion a n d nitrif ication/denitrification). Inpu t minus total rernoval processes; that is. the output load expected based on the measured removal processes. The actual output load rneasured.

i The difference between the calculated output and actual output. Positive values indicate al1 removal processes were not measured (Le., treatment was underestimated). Xegative values indicate rernoval processes measured overestimated treatment.

nm not measured

- - fond D y k e

Ammonia Input Load (g/m'/season)

Figure 2.1 Ammonia removal (%) for Dyke (1992 - 94) and Pond (1992 - 93) water.

Figure 2.2 Zitrite-nitrate nitrogen concentration in t h e Dylie water replicate t renches at Stations A. B. C and D. 1994.

Staûon P

astabon B *Station C

0-D

Figure 2.3 Ammonia-ammonium nitrogen in the Dy ktt hater replicatc- trenches at Stations A, £3, C and D, 1994. (a) concentration. (b ) as a % of concentration at Station A

k-igure 2.4 Exchangeable ammonia nitrogen (mean 2 S D ) in t h e sediments of the constructed wetlands trenches. 1993 a n d 1994: b y date.

k ' i 1 ri Jlid Out I ri \ltd Out

Figure 2.5 Exchangeable arnmonia nitrogen (mean 2 S D ) in the sediments of the constructed wetlands trenches, 1993 and 1994: by zone.

Figure 2.6

ncl

data no iata

no data

Pond

O Cana! l

BulwsI4

no data

Pond

Total nitrogen in the storage pool in 1993 and 1994 at t h e end of the growing season (Septem ber ) .

J une lut>

Figure 2.7 In s i tu denitrification (mean I SD) measured us ing the acety lene blockage technique in Control and Dy ke Drainage trenches. 1994.

In h.1 id Out

Figure 2.8 Arnmonia volatilization rates (mean 2 SD) from the ConstrucLcd Wetiands trenches, 1994.

(a) Control Output to amosphere t ) 02 mg munonia-\ m : d' l i 114 mir nitrous osidc-\ ni - d- i,

Figure 2.9

To Sediments (37.4%) 1 0 1 mg N m-'d-' Output vta water ( 21 3' o i

,4 mg ammontri-S m d 1 7 mg nitnic-N m-' d

The nitrogen budget in the (a) Control a n d ( b ) Ilyke frenclies in 1994. Boxed quantities arp mass stored ( m g m-'): othei- quantities are fluxes (mg m d ) : number in brackets is percent of i n p u t nitrogen.

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3.0 HYDROCARBON REtM0VA.L IN CONSTRUCTED WETLANDS RECEIVTNG OIL SANDS WASTEWATER

3. i introduction

Significant sources of ene rgy a r e cur ren t ly obtained from mining of oil- rich s a n d s in nor thern Alberta (cur ren t ly more than 13 % of Canada's annuai oil needs). Active mining leases dur ing this s tudy w e r e about 26,600 ha, which produced about 300,000 barrels of oii per day (bpd). Recent approvals for new mines include development of another 18,400 ha. Should al1 mines tha t are cu r ren t ly proposed becorne act ive more than 49,000 ha would be disturbed. These su r f ace mining processes d i s tu rb large areas of the landscape which will r equ i r e reclamation. Additionally, the extraction process p r o d u œ s la rge volumes of waste b j l i ngs t ha t mnta in various hydrocarbons and ammonia. This s tudy provides a pilot experiment using constructed wetlands to mitigate the environmental impacts of oil s a n d s mining.

Suncor Inc. Oil Sands Croup (Suncor) is mining oil sands in nor theastern Alberta, Canada, 2 km north of Fort XcYurray. The mining and

- subsequent conversion of oil sands bitumen to synthetic c r u d e oil began in 1967. The Clark Hot Water Extraction Process used t o e s t r ac t bitumen from t h e oil sands produces a waste tailings. These tailings. made up of sand. clay, water and unrecovered bitumen, are stored in large tailings ponds (Figure 1.2). One cornponent of tailings. the fine clays known as f ine taiis, dewater very slowly. Therefore, there is potential for the release of contarninated water as leachate and/or sur face run-off to te r res t r ia i and/or aq uatic systems as these tailings ponds a r e reclaimed.

A treatment system will be req uired t o mitigate any environmental impact of leachate and/or run-off water. Both organic (e-g.. hydrocarbons) and inorganic ( e.g., ammonia) contaminants are of concern. A wetlands treatment system was investigated for i t s potential for sustainabii i ty over t h e long term and subsequent low cost. Leachate from sa tura ted tailings pond sand dykes (hereaf te r re fe r red to as Dyke water) and tailings pond top water (hereafter re fe r red to as Pond water) were treated in a pilot f ield-scale facility .

A variety of physical, chernical and biological processes within wetlands are responsible for removing contaminants. However, there are only a fen;

examples of wetlands being used specifically to t rea t wastewaters from petrochemical facilities or o ther sources of organic chem ical wastes (Hamilton et al., 1993). The primary contaminants in this has tewater are naphthenic acids and ammonia. However, t h e focus of th i s p a p e r is naphthenic acid removal.

Naphthenic ac ids or iginate from petroleum c rude oil and oil s a n d s bitumen r e se rves (Fan, 1991: AEP. 1996). Yono- and poly-cycloalkane (cyclohexane and cyclopentane) carboxylic acids with aliphatic s ide chains of various Lengths predominate in this group. Naphthenic acids with 2 r ings predominate oil s a n d s tailings (Lai et al,, 1995). Naphthenic acids have propert ies associated with sur fac tan ts and this is thought to b e one of t h e fac tors contr ibut ing to t h e low settling rates of oil s ands tailings (XEP,

1996).

Naphthenic acids are of concern because they are tosic and corrosive (Liorales et al.. 1993). Compounds in t h e acid fraction of < 1000 molecular weight precipitation a t pH of 2.3 account for 80 - 90 % of t h e total acid fraction which accounts for 5.5 - 100 % of t h e toxicity observed in oil s a n d s

wastewater ( YacKay and Verbeek, 1993). This fraction has been identified as naphthenic acids (Liorales et al.. 1993).

Saphthenic ac ids are difficul t to analyze b j - conven tional mass spectrometr ic (YS) techniques because of their polaritl- and increasing non-volatility with increasing molecular weight tech niques ( ' lorales et al.. 1993). Although the re have been improvernents (such as Fast Atom

Bombardment 'iass Spectrometry). these methods are still not ideal for quant i ta t ive analyses of cornplex mixtures ( Liorales et al,, WU). Fourier Transform Irifrared (FTIR) Spectroscopy has b e n rnainly used fo r quali tat ive analysis al though recently usage for quantitat ive analysis h a s increased. In fact, a quanti tat ive method was developed fo r measuring naphthenic acids by Syncrude Canada Ltd.; however. this w a s not available unti l 1995.

In this s t u d y , a method was required tha t could quantify the naphthenic ac ids input to and ou tpu t from t h e wetlands. As YS techniques utility w e r e Iimited and expensive and t h e FTIR method w a s not avaifable, a s u r r o g a t e for naphthenic acids that could be quantified was required. Total extractable hydrocarbons (TEH) were analyzed as a s u r r o g a t e for naphthenic acids. The chromatograph of commercial grade naphthenic acids

a n d oif s a n d s wastewater were similar and exhi bit t he same characterist ic broad peak between C E and C28 (Figure 3.1).

A varie ty of physical, chemical and biological processes in wetlands rnay ac t to translocate and transform hydrocar bons- In wetlands receiving hydrocar bons, these processes may act together to effectively rernove hydrocarbons from t h e water column (Le.. retain hydrocarbons such that ou tf low water q uality is im proved ). Physical, chemical and biological translocation processes inchde: particulate settling, diffusion of dissolved forms, sorption of soluble forms on substrate. volatiiization and plant uptake. I t is important to realize that t hese processes have opposite/countervaiiing processes which may r e t u r n hydrocarbons t o t h e water column (e-g , , resuspension, Iitter fall, diffusion). There are aiso transformation processes tha t may remove hydrocar bons ( e.g,. mineraiization via microbial degradation) from the water column (Kadlec and Knight, 1996).

I t i s important to unders tand the translocation and transformation processes acting to remove chernicals of environmental concern when designing treatment wetlands systems (Kadlec and Knight, 1996). That is. are t h e contaminants simply retained in the wetlands (e-g. , sequestered in sediments o r plants) and potentially available t o r e tu rn to t h e water column. ra ther than removed from the system ( e.g., photolysis, volatilization. mineralization ). If contaminan ts are not removed they may be released back into the water cohmn later, therefore reducing wetland treatment. The expected removal mechanisms a r e microbial transformation a n d sedimentation translocation processes: althoug h plant uptake rnay ais0 cont r ibu te to hydrocarbon removal i t is not expected to be significant.

Indigenous microbial communities have degraded naphthenic acids (Herman et al., 1994; Lai et al., 1996). Nutrients and temperature were shown to affect biodegradation rates (Lai et al., 1996). Wetlands provide more sui table habitat for microbid comrnunities compared with tailings ponds because of increased sur face area and nut r ien t r ich resources: consequently, biodegradation i s expected to occur in wetlands.

Other removal mechanisms inciude volatilization. sedimentation and plant uptake. Saphthenic acids are non-volatile (Morales et al.. 1993) and therefore photolysis and volatilization i s probably not a significant removal pathway ( AEP, 1996). Sedimentation with solids i s possible, while organics may also be b k e n u p by plants (Sirnonich and Hites, 1995).

- area - - water volume -

X

- liner - subs t r a t e -

175 m' 3 52.5 m (based on 50 m x 3-5 rn x 0.35 m

0.9 porosity) 40 mil high density potyethylene 0.45 m of sand topped with 0.15 m of

orgânic muskeg soi1 - vegeta tion - 300 Typha latifofia (cattail) shoots and 60

Scirpus validus (bu l rush ) culrns ( bearing 6 - 15 shoots) were planted in each t rench in J u n e 1991

establishment - 1 year

Experimentai Design

Treatment waters were applied to t h e nine t renches in a randomized block design. Three replicate Control trenches were corn pared with treatment t renches (Dyke or Pond). T h e experimental design changed (Le., flow rates) from year t o year (Table 2.1).

Yonitoring Programme

There w e r e two sampling schemes depending on t h e type of analysis: 1)

four sampling s ta t ions from inflows to outfIows (Le., A, B, C. D) w a s followed for water quality and water chemistry. and 2 ) th ree stations from inflows to outflow (Le., in, mid, ou t ) was followed for sediment chemistry. For t he four station sampling scheme. Station A sarnples were taken from the inflow pipes , Station B was located 15 - 17 metres downstream of the inflow to each t rench, Station C 30 - 34 metres downstream of t h e inflow. and Station D sarnples were taken from the outflow pipe/weir (50 rn downstream of the inflow). For t h e th ree station sampling scheme, ''in" sarnples were collected from within t h e f i r s t third ( O - 15 m) of each t rench, t h e "mid" from the second th i rd (16 - 30 m) and the "out" from t h e last third (31 - 45 m).

Sarnpie frequency was limited in 1992 due to water limitations (i.e., water was transported by truck in 1992 r a t h e r than pumped as in 1993 and 1994). In 1993 and 1994, water chemistry was sampled as frequent ly as weekly to ensure that t h e sample frequency w a s at least as often as trench turnover fi.e., within t h e estimatecl mean hydraulic retention time of 7 to 18 d a y s ) . Sediment sample frequency w a s once per study year.

3.22 Sample Collection

Kater samples were collected directly into sample bottles a n d sh ipped to the laboratory (ASL in Vancouver, BC) t h e same day such that sample analysis was initiated within a seven day holding time.

Sedimen t

Sediment samples were collected using a plastic mrer ( 5 c m in diameter) with only t h e top 1 c m being collected. Several grab samples were collected at each station and placed in "whirl pack bags" to ensure a minimum of 100 g of sample.

3-2.3 Sam ple Analysis

The analysis of Total Estracta ble Hydrocar bons (TEH) bas conducted in accordance with CSEPA Liethod 3510/8013 (CSEP.4. 1984). The procedure involved the extraction with methylene chloride. The extract w a s then reduced in volume and analyzed by capillary column gas chromatography w ith flarne ionization.

3.2.4 TEH Budget

TEH were measured in sedirnents during 1993 and 1994. Yass balance equations were determined by estirnating the inputs. ou tpu t s and sedirnent s torage.

The following mass balance equation can b e applied to processes determining t h e fate of contaminants within wetlands:

Depending on the compartments and processes measured. this equation can be rewrit ten to estimate the unknow n compartment/process.

3.2.5.1 W a t e r Chemistry

T h e water chemistry da ta were not stat is t ical ly analyzed because of l o ~

sample numbers at inflow s ta t ions (Le., composites were taken from the t h r e e replicate t r enches to reduce cos t s ) a n d a lack of variat ion (as rnany Control sarnples r e su t t s were less than the detection limits), Data were graphically in te rpre ted t o assess w he the r t h e r e was a dif ference between treatment waters or from inflow to outflow for a given wastewater.

3.2.5.2 Sediment Chemistry

Sediment concentra t ion (mg kg-') da ta were tes ted fo r significant difference amongst treatment waters. Al1 stat is t ical analyses were performed using t he SYSTAT sta t is t ica l package (Wilkinson. 1990). A probability of a s 0.05 was used to def ine statistical significance.

Repeated measures ana lys i s of variance (-4XOVA) tests were conducted on t he TEH concentra t ions in sediments t o assess t h e ef fects of wastewater type. Individual trenches were used as replicates of t h e treatment water t ype (i.e., t h e t rea tments were sampled in tr ipl icate) . Yeasurernents were replicated three times in each sampling period (i.e., d a t e ) or zone (Le-. in.

mid. o u t ) for most sediment data. Repeated measures des ign is the same as a spl i t plot design. with t h e t rea tments applied to t h e whole plots o r t renches and t he sampling period and zone appiied within plots or t renches (Le.. t h e t renches were repIicates fo r t e s t ing treatments, bu t blocks for tes t ing sampling period or zone effects) . '(:O t r ench effects were assurned. Post- hoc con t r a s t s were used t o determine w hich means were significantly different us ing Bonferroni ad jus ted a lpha levels (Le.. cri t ical level of 0.05 divided by the number of con t ras t s t e s ted) . Testing a subse t of corn parisons r a t h e r t han ai l possible pairwise comparisons ( e.g., Tukey H S D ) often gives more power to the analys is (Wilkinson. 1990). Repeated measures ANOVAs were conducted for each parameter (e.g., total nitrogen in bulrush root. exchangeable ammonia in sediments. etc.) and year separate ly (Le., 1993 and 1994).

w ith th is s tudy. Direct concentration cornparisons are not a p propriate, b u t a cornparison of rernoval efficiency i s useful. Oil and grease removal efficiency reported in t h e l i t e ra ture ranged between 53 % to 96 % (Table 3.3). This efficiency was higher than t h e efficiency observed in th i s s tudy . However. input loading rates fo r t h e l i t e ra ture s tud ies were lower than in this s t udy , sugges t i ng tha t t h e lower removal efficiencies reported in th i s s t udy were d u e to t h e h igher input loading rates.

3.3.3 Sediment Chemîstry

Total extractable hydrocarbon concentrations were lower in the sediments of the Control t renches than e i ther the Dyke and Pond t renches except in 1993 for Dyke t r enches (Table 3.4). However, t h e s e t r e n d s were not significantly di f ferent in a n y year. No significant t r e n d s iri the sediments frorn inflow to outflow were observed.

3-3.4 TEH Mass Balance

The mass balance for TEH in t h e constructed wetlands treatment t renches was determined for each s t u d y yea r (Table 3.3). TEH input and ou tpu t loading and change in the sediment s to rage cornpartment were rneasured. and frorn this information a mass balance in Dyke and Pond t renches was calculated. Other processes t ha t could have con t r ibuted to hydrocarbon removal include volatilization. plant uptake and microbial mineralization. These processes were no t measured in this s tudy .

Removal to the sediments accounted for 8 - 27 % of the TEH retained. The unaccounted TEH load could have b e n volatilized, taken u p by plants o r mineralized by microbes. Volatilization was not expected to be signif icant

because naphthenic ac ids were expected to make u p t h e majority of TEH and these compounds are not readily volatilized (Morales et al., 1993). Althou g h plant u ptake and microbial mineralization were not measured in

t h i s study. parallel studies were conducted in t he se same constructed wetlands t renches to assess t he se potential removal rnechanisms (Nix et al., 1994 and 1993).

The paraIIel s tud ies found Little or no TEH uptake by t h e emergent aquat ic macrophytes (hereaf ter r e f e r r ed to as macrophytes) which make u p the rnajority of the plant biomass ( N i x et al. 1994 and 1993). In t h e parallel

studies, carbon dioxide (CO2) production w a s measured as a n indicator of

mineralization. The final product of complete biodegradation (i.e., rnineralization) of hydrocarbons i s CO,. Six et al. (1994 and 1995) modified a method used to measure COt production ra tes from soils (Anderson. 1982).

In 1994, about 6.8 kg C/wetlands/season (as C O I ) w a s measured leaving the Dyke trenches; however, only 1.9 k g C (or 2.2 k g TEH assuming carbon made u p 85 %) were retained by the wetlands (Le., t h e inflow load minus t h e outflow load). This sugges ts tha t t h e CO2 evolved from t h e trenches represents other processes in addition to mineraüzation of h y drocar bons.

The amount of CO, evolved from t h e Dyke and Pond trenches w a s corrected by CO2 evolved from or respired in t h e Control trenches, t o account for any processes other than TEH mineralization. However, t he difference d u e to nitrification/denitrification would not have been included in this "correction" as elevated ammonia only existed in Dyke and Pond trenches. In 1994, about 3.6 kg ammonia -i were retained by t he t renches (not including what was retained by the sediments o r macrophytes), which would produce 4.6 k g C assuming the unaccounted for ammonia was nit rif ied /den i t r s i ed . S u b t r a c t i n g th i s ca rbon d u e to nitrification/denitrification from t h e total evolved means about 2.2 kg C was

evolved due to mineralization which is about equal to the unaccounted for carbon retained (i.e.. 1.9 k g C ) by the trenches. This. means t ha t the corrected CO, evolved from Dyke and Pond trenches iras close to the s u m of CO2 produced from nitrification/denitrif ication and TEH mineralization.

In 1993, however, t he re were much Iower CO, evolution rates (Six et al., .I

1994). lviineralization did not seern to contribute s u bstantially to hydrocarbon retention by the t renches once the above consideration for nitrification/denitrification was included. I t seems unLikely that the difference could be made u p by losses via volatilization. T h e primary removal mechanisrns were likely rnineralization and sedimentation. but t h e

contributions from each could not be calculated from the available data.

3.4 Conclusions

Average TEH removal rates (1.e.. mass reductions) ranged from 19 % to 76 % for Dyke water and 54 % to 69 % for Pond water during t h e th ree year s tudy. These rates were lower t h a n rates for oil and grease reported in the l i terature (Table 3.3). The difference w a s probably due to t h e amount of contaminant loading. Input loading was on average higher in this s t u d y and ranged from 146 to 1143 mg m-2 d-' compared with 32 to 168 mg d-l

for oil and grease in the other studies. TEH rernoval rates varied by year. Year 1 (1992) had a relatively low loading rate which resulted in a high treatment efficiency. Year 2 (1993) had a very high loading r a t e which resulted in lower treatment efficiency than in 1992. Year 3 (1994) had an intermediate Loading rate which resulted in an intermediate loading rate for Dyke water.

TEH removd to the sediments accounted for less than 30 % of t h e TEH load

retained by t h e wetlands in this study. Other rernoval mechanisms include volatilization, plant u ptake and micro bial mineralization, Volatilization was not expected to b e significant and parallel s tudies did not lndicate plant uptake was significant and the removal via mineralization was unclear. Further method development is required to e n s u r e removal processes and s torage are reflected accurately. Aithough the fate of TEH remained unclear, removai w a s substantial and warrants f u r t h e r investigation to assess the role of constructed wetlands in the oil sands industry with respect to hydrocarbon treatment and reclamation.

T a 3 . Total extractable hydrocarbon in water (mg L-') for the inflow (A ) and outflow ( D ) of the constructed wetlands t renches.

-- - - . - - - - -- - - . . . - - -

Year Treatment Station Yin Yax Y ea n SD Water

Control

Dyke

Pond

Control

D y k e

Pond

Control

Dyke

a samples were collected 4 times in 1992: t rench inflows w e r e composited (n=4) and outflows w e r e individual samples (n=12)

b sarnples were collected 6 times in 1993; trench inflows were composited (n=6) and outflows w e r e individual samples ( n= 12)

c samples w e r e c o l l e c t e d 9 times in 1994; trench i n f l o w s were cornposited (n=9) and outflows were individual samples (n=27)

Table 3.2 Total extractable hydrocar bons (TEH) removal in constructed wetlands receiving oil sands wastewater.

Year/ TEH Inputs (mean TEH Outputs Removal Load Water ,+ SD) (mean 2 S D ) ( % of Removed

inflow) (mg rn-'d-') mg L*' mg r i 2 d-' mg L-' m g m-' d-'

1992 D y k e 13.8 220 4 .O 52 76 168

Pond 12.2 211 7.9 65 69 146

1993 D y k e 20.4 992 20.8 799 19 193 Pond 46 2116 25.7 973 54 1143

1994 Dyke 37.6 696 27.5 484 30 212

Table 3.3 Oil and grease removai in various wetlands treatment facilities.

Input ~oad' Inflow Outflow Removal Reference ( m g rn-' d) m g L-' mg L-' (%)

32 2.1 0.130 94 Litchfield, 1993

168 0.840 0.290 63 Dong and Lin. 1994

- 12.6 0.490 96 Ramirez. 1993

113 4.3 2.0 53 Duda. 1992

a calculated from information given: information to calculate output loading was not given

- no information given to calculate input loading

Table 3.4 Total extractable hydrocarbons in the sediments (mg kg-!) in the constructed wetlands trenches.

Year Treat ment Mater Yean SD n

1992 Control

Dyke

Pond

1993 Control

Dy k e

Pond

1994 Control

Dyke

Table 3.5 Total Extractable Hydrocarbon mass balance for Dyke and Pond water.

a- . a after 90 days of wastewater application: : iig r r e n c d season" = 5 ; ie a'- n a - : ,

after 116 days of wastewater application: ! I:e trench-' seasorÏi = 13 a? i*' -, i*' i

af te r 101 days of wastewater application: 1 icg irencs-' season-: = 37 ay a-- d m ' &ter 133 days of wastewater application: 1 kg trench-i reasoo-' = 43 16 a-' d - ' ' The surn of aii rernoval to s torage pool (sediments and uptake b y p l a n t s ) a n d by p r o c e s s e s ( v o l a t i l izat ion a n d nitrification/denitrification ) . ' Input minus totai removal processes: tha t is. the output load expected based on the measured removal processes. The actual ou tput load measured. ' The différence betueen the caiculated output and actual output. Positive values indicate al1 removal processes were not measured (i.e., treatment w a s underestimated ).

nm not measured

Parameter

Input (loading) k g TEH/wetlands/season

Total Removal roce es ses^ Sediment Storage

k g TEH/wetlands/season 1

Calculated Output (Input - Removal) f

k g TEH/wetlands/season

Actual Output (untreated load in outf10w)~

k g TEH/wetlands/season

~ y - k c "

3.5

nm

-

0.8

2.2

ponda

3.3

nm

-

1 .O

Dykr R

16.2

2.7

13.3

11.3

16.7

~ ~ k t ? ~

20.1

2.5

17.6

1 6

Iinderestirnated treatment performance h

k g TEH/wetlands/season

pondC

37.4

3.3

33.9

17.2

- - 1.4

Naphthenic Acid Standard

inflow -

I

I

Pond 1A - oufflow - A

1 1 I 1 1 1 1

5 IO 15 20 25 30 35 7me (min)

Figure 3.1 A cornparison of GC chromatograrns for a naphthenic acid standard and Pond inflow and outflo* water. (Nix et al.. 1993)

1 O 0 150 t ( ,O

Hydroearbon Input Load (g/m2/seas<in )

Figure 3.2 Hydrocarbon removal (%) for Dyke ( 1992 - 94) and Pond \rater (1992 - 93).

3.5 Literature Cited

Alberta Environmental Protection ( AEP). 1996. Saphthenic acids background information: Discussion Report. Environmental Criteria Branch. Environmental Assessment Branch. Environmental Regulatory Services, Alberta Environmental Protection. Draft Repor t . February 29, 1996. 79 pp.

.Anderson, J.P.E. 1982. Soi1 respiration. In: A . . Page (ed, ). Yethods of Soi1 Xnalysis: P a r t 2 - Chernical and Liicrobioiogical

Properties. Yadison. KI. pp. 383-845

P H . 1989. Standard Yethods for the Examination of hater and ivastewater, 17th Ed. American Public Health Association. Mater Pollution Control Federation and American Wastewater -4ssociation. Washington D.C.

Dong, K. and C. Lin. 1994. Treatment of petrochernical w a s t e ~ a t e r by t h e system of wetlands and oxidations ponds and engineering design of the system. Preprint of paper presented at the 4th International Conference on Wetlands for Water Pollution Control, Guangzhou. PRC. (cited in Kadlec and Knight, 1996).

Prepared Califor nia Richmond,

Fan, T.-P. 199 fast atom

Duda, P.J. 1992. Chevron's Richmond refinery water enhancement wetland. for The Regional hater Quality Control Board, Oakland by Chevron CSA Inc,. Environmentai Operations Division.

, California. Yarch 1992-

1. Characterization of naphthenic acids i n petroleum by

bom bardrnent mass spectrometry. Energj- F u e l s 5:371-375.

Hamilton, S.H.. P.C. Nix and -4.B. Sobolewski . 1993. An overview of constructed wetlands as alternatives to conventional waste treatment systems. tuater Pollution Research Journal of Canada 28(3):529-548.

Herman. D.C., P.M. Fedorak, Y.D. MacKinnon and J.h. Costerton. 1994. biodegradation of naphthenic acids by microhial populations indigenous to oil sands taiiings. Canadian Journal of .Wcrobiolog,r. 40:467-477.

KadIec, R,H, and R.L. Knight, 1996. Treatment kleflands. Boca Raton: L e w i s Publishers. 893 pp.

Lai, J.W.S., L.J. Pinto, E. Kiehlrnann, L.I. Bendelly-Young and Llargo '1.

Lioore. 1996. Factors that affect the degradation of naphthenic acids in oiI sands wastewater by indigenous microbial comrnunities- Environmental Tox'coIogy and Chemis try 15(9):1482-1491.

Litchfield. D.K. 1993. Constructed wetlands for wastewater treatment at Amoco Oii Company's Yandan, 'r'orth Dakota, refinery. In: G A Moshiri (Ed. ). constructed Wetlands for water quaiity improvement. Chelsea, ?II: Lewis Publishers. pp. 485-488.

YacKay. K.C. and A G . Verbeek. 1993. The characterization of acutely toxic compounds in fine tails pore water and tailings pond water. Final Report. Prepared for Alberta Department of Energy, OiI Sands and Research Division (AOSTRA #8878Y) by Lniversity of Alberta. 47

PP*

Yorales, -4.. S.E Hrudey and P A , Fedorak. 1993. Yass spectrometric characterization of naph then ic acids in oil s a n d s xastewaters: analysis, biodegradation and environmental significance. Summary of

Final Report. Prepared for T h e Fine Tailings Ftindamentals Consortium by Cniversity of Alberta. 92 pp.

Nix, P.C., S.H. Hamilton, E.D. Bauer and C.P. Gunter. 1993. Constructed wetlands for the treatment of oil sands wastewater: Technical report

#2. Prepared for Suncor Inc., Oil Sands Group by EVS Consultants, North Vancouver, B.C. 73 pp. + tables, figures and appendices.

N i . . P.C.. S.H. Hamilton. L. Bendell-Young, C.P. Gunter, F.S. Bishay and Y.D. Paine. 1994, Constructed wetlands for the treatment of oil sands wastewater: Technical report #3. Prepared for Suncor Inc., Oil Sands Group by EVS Consultants. North Vancouver. B.C. 223 pp. +

appendices.

Nix. P.G.. F.S. Bishay, S.H. Hamilton and Y.D. Paine. 1995. Constructed wetlands for the treatment of oil sands wastebater: Technical report

#4. Prepared for Suncor Inc., Oil Sands Group b y EVS Consultants.

North Vancouver, B.C. 369 pp. + appendices.

Ramirez, P. 1993. Contaminants in oil field produced waters discharged into Lock Katrine WetIand Cornplex. Park County, t\; yoming and thei r bioeoncentration in the aquat ic bird food chah. Prepared for L S Bureau of Land Lianagement, Cody Resource Area. Cody, hyoming by Ecological Services Office, Cheyenne, Wyoming. L'S Fish and kildlife Service, Region 6, Environmental Contarninants Program. 37 pp.

Sirnonich, S.I. and R.A. Hites. 1995. Organic pollutant accumulation in vegetation. En vironmental Science and Technology 29( l2):2905-29 13.

Cnites States Environmental Protection Agency (CSEPA). 1984.

Guidelines Establishing Test P rocedures fo r t h e Analyses of Pollutants unde r t h e Clean Water Act. 40 CFR Part 126- Federal Register 49:209.

Kilkinson, L. 1990. SYSTAT: the system fo r s ta t i s t ics . SYSTAT. Inc. Evanston, Illinois. 677 pp.

4.0 EMERGENT AQU ATIC MACROP HYTE P RODUCT ION AN B DECOMPOSITION IN CONSTRUCTED WETLANDS RIECEIVING OIL SANDS WASTEWATER

4.1 1 n trod uction

Significant sources of ene rgy are current ly obtained from mining of oil- rich sands in northern Alberta (current ly more than 15 % of Canada's annual oil needs). Active mining leases during this s tudy w e r e about 26,600 ha, which produced about 300,000 barrels of oil per day (bpd). Recent approvals for new mines include development of another 18,4GO ha. Should al1 mines that are current ly proposed becorne active more than 49.000 ha would be disturbed. These surface mining processes dis turb large areas of the landscape which will require reclarnation. Additionally, the extraction process produces large volumes of waste tailings that contain various hydrocarbons and ammonia. This s tudy provides a pilot experiment using constructed wetlands to mitigate the environmental impacts of oil sands mining.

Suncor Inc., OSG is current ly mining oil sands in north eastern Alberta, Canada, 23 km north of Fort YcMurray. T h e biturnen (i.e., oii) is extracted from the oil sands by the Clark Hot Water Extraction process (Gulley and Klym, 1992). Fine tailings, a hy-product of the process, is made u p of clay, water and unprocessed bitumen, and is stored in tailings ponds on the Suncor lease. Wastewater associated with these ponds must be detoxified as part of t h e reclamation scheme(s) for the Suncor site. Preliminary evidence suppor ts t h e potential use of wetlands to treat some of these wastewaters. A wetland that established accidentally in d y k e seepage from a tailings pond reduced contaminant concentration and toxicity from inflow to outflow (Hamilton and Nix, 1992).

The ecological consequences of using constructed wetlands to treat oil sands wastewater were investigated. Aquatic macrophytes (hereaf'ter referred to as macrophytes) have direct and indirect effects on contaminant removal and treatment efficiency, as wel l a s being potential indicators of ecosystem health. Macrophytes can remove contaminants b y

direct uptake, and/or provide the necessary environment for physical removal mechanisms and habitat for microbial removal mechanisms. Yacrophyte production and decomposition play direct and indirect roles in improving water quality. and therefore they were studied to understand

the viability and sustainability of constructed wetlands receiving oil sands wastewater.

Yacrophytes have the ability to directly remove contaminants by metabolic and absorption processes. Phn t metabolism (both biosynthesis and catabolic reactions ) of ref ractory organics. bacteria and virus: and plant absorption of nitrogen. phosphorus, met& and refractory crganics are direct removal processes which may occur in wetlands treatment systems (Watson et al., 1989). Macrophytes have been shown to take u p metals (e.g.. Kadlec and Tilton. 1979: Yorea et al., 1990; hmd and YcAtamney, 1994.

etc) and nutrients (e-g., Gersberg et al., 1984, 1987: Nichois, 1983; Rogers et al,, 1991; Tanner et ai., 1995). The role of plants in metal rernovd is secondary to sedimentation (Dunbabin and Bowmer , 1992), and their role in nu t r i en t rernovai is often secondary to rnicrobial and sedimentary processes. However. under specific conditions nutrient u p b k e by plants may be t h e primary removal mechanism (e.g., Rogers et al., 1991). Although the contribution of plant uptake to overall removal is generally Low. macrophytes are important because they enhance removal by other p rocesses.

Yacrophytes indirectly affect contaminant removal in wetiands receiving wastewater. Yacrophytes stabilize sediments with their roots and rhizomes, which prevents erosion (Dunbabin and Bow mer, 1992). They also disperse

floclv, further reducing the threat of erosion and channelization. Conmmitantly, sedimentation rates are increased as water velocity is reduced (Dun babin and Bowmer , 1992). Yacrophytes (live and dead biomass and leaf litter) provide a surface for degrading/mineralizing bacteria (Le., biofilrn). Yacrophytes also increase the oxygen transfer to the sediments by oxidizing the rhizosphere. This oxygen transfer supports aerobic microbial populations in the rhizosp here (Brix, 1990). These microbes modify nu trients, metals and trace organics which of ten results in their removal from the wastewater (Gersberg e t al., 1986). Nitrogen can be removed to the atmosphere by nitrifying/denitrifying bacteria ( H a m r n e r and Knight 1994). Y e t a l s precipitate to the sediments with hydrogen sulphide. produced by sulphate reducing bacteria (Faulkner and Richardson, 1989). while organics are rnineralized by bacteria. These modifying microbes require nutrients which can be obtained through the decomposition of rnacrophytes which recycles nutrients back into the system. Regardless of whether macrophytes play a direct or indirect role in contaminant removal their contributions to wastewater treatment cannot be overlooked.

Production and decomposition of macrop hytes a r e important processes in a healthy wetland. In o rde r for the macrophyte community to provide habitat for bacteria and other organisms, take up nutrients and oxygenate t h e sediments, healthy, viable vegetation must be maintained - Decornposition of annual dieback vegetation must also be maintained to recycle nutrients back into the system thereby providing the necessary resources for the microbiai community that degrades contaminants. Hence macrophyte production and decomposition was used as a measure of wetland sustainabiiity in this study.

The objective of this s tudy w a s to determine the deleterious impacts of oii sands wastewater on macrophyte production and decomposition in cons luc ted wetlands. The nul1 hypotheses tested were: 1 ) no difference in macrophyte production in constructed wetlands trenches receiving different test waters, and 2) no difference in macrophyte decomposition in constructed wetlands trenches receiving different test waters. Nitrogen dynamics and nitrogen and hydrocarbon retention b y t he constructed wetlands were addressed in separate chapters (Chapters 2 and 3, respectivefy ).

4.2.1 Site Description and Experirnentai Design

The Suncor experimental constructed wetlands are located in northeastern Aberta, Canada, 25 km north of Fort YcYurray (3'1"04*N 111°28'\Y). They are located southwest of Tailings Pond 2 between the tailings pond and Highway 63 (Figure 1.1). They were constructed in Yay and J u n e 1991 and cover an area of approximateiy 1.4 ha (Figure 1.3). Nine trenches were constructed in 199 1 with the follow ing characteristics:

cell dimensions - 50 m by 10 rn a t the top and 2 m at t h e bottom

side d o p e - 2:l trench slope - 0.5 % (equivalent to a drop of 0.25 m over

50 in)

mean depth area

water volume

liner subs t ra te

vegetation

3 - 52.5 m (based o n 50 m x 3.5 m x 0.35 rn x 0.9 porosity) - 4 0 mil high densi ty polyethylene - 0.45 m of sand topped hi th 0.15 rn of

organic rnuskeg soi1 - 300 Typha latifolia (cattail) shoots and 60 Sckpus validus (bu l rush ) culms (bearing 6-

13 shoots) were planted in each trench in June 1991

establishment - 1 year

The trenches were planted with Typha latifdia ( 10 cattail shoots per 6 rnZ)

and Scripus sp . (2 bulrush culms of 6 - 15 shoots each per 6 ml) in early June 1991, and allowed to establish for the remainder of the 1991 growing season. These planting densities were similar to those used in o ther s tudies (e.g., Cooper and Hobson. 1989; Van Oostrom and Cooper, 1990; Pullin and H a m m e r . 1991). Typha latifolia was selected because Typha spp.

are hardy. widespread, easy to grow. and are capable of surviving under adverse environmental conditions (Desjardins e t al., 1986: YcNaughton et al.. 1974). Scirpus vaiidus w a s selected due to its super ior ( to Typha spp.) ability t o t ranslocate oxygen to the roots (Gersberg et al.. 1986). In addition. both plants were selected because they inhabited the area prior t o wetland construction. Donor s i tes ( w i t h limited impacts from t h e industrial s i te) were easily Iocated on the Suncor mine lease.

Experimen fa1 Design

Treatment waters were applied to t h e nine t renches in a randomized block design. Three replicate Control t renches were compared wi th treatment trenches (Dyke o r Pond). The experimental design chariged (Le.. flow ra tes ) from year to year (Table 2.1).

In addition to t h e constructed wetland trenches, a n adjacent contaminated wetland (called t h e Natural Wetland) and four reference wetlands were monitored. The Natural Wetland was located adjacent to the constructed wetlands system and received unknown amounts of Dyke Drainage water from Pond 2 as well as inputs from sur face runoff water. Over the t h ree s tudy years four different reference wetlands were monitored (Figure 4.1).

Together these wetlands provided information on or about longer term impacts and natural variability.

.Vonitoring Programme

Three stations, pe r trench, from inflow to outflow (Le.. in , mid, o u t ) were sampled. The "in" sarnples were collected from within the f i rs t third ( O - 15 m ) of each trench, the "mid" from the second third (16 - 30 m) and the "out" from the last third (31 - 45 m) . Sample f r e q u e n c y varied depending

on the s tudy year and is described below.

4.2.3 Production

In 1992, trenches w e r e sampled only in August, while in 1993 and 1994 they were sampled in July, August and September. N e t primary production w a s estimated by measuring peak above-ground standing crop. The experimental design for the production measurements varied sornew hat year to year (Table 4.1). Three to five 0.25 or 1 m L quadrat plots p e r zone were harvested. In 1992. 1 m' quadrat plots w e r e used. b u t later as vegetation became more dense, quadrat number and size were reduced to three 0.25 m2 in 1993 and 1994.

Live above-ground tissues were clipped at the soi1 surface in each plot. and plants were sorted into three groups for Iater drying and weighing: cattaii, bulrush and al1 other species. Khen possible, plants were identified using Moss (1983). Niering (1989) and Weinmann et al. (1984).

Below-ground biomass w a s sampled on ly in the September survey in 1993 and 1994. In 1993, a coffee m n (15 c m diameter; 17 c m depth) was used as a corer. One end of the can had been removed while several holes had been put into the other end. Once the above-ground vegetation had been removed the corer w a s piaced in the centre of t h e plot and pushed into the sedirnent using a spade to separate t h e corer and i ts contents from t h e surrounding sediment and roots. The cornpiete recovered core was stored under cool conditions until sieved. Upon sieving. biomass w a s sorted into current years and older growth for both cattail and bulrush, a n d other biomass was lumped together. New rhizome growth w a s relatively easy to differentiate from older growth, unless rhizomes were discoloured by the anoxic sediments, at which t i m e differentiation became somew h a t subjective. Wei weights were measured and the samples were then dried a t '40°C and weighed again. Since results were highly variable in 1993, in 1994 a less labour intensive samplîng method was used. Single shoots were collected

(above- and below- ground parts). Above- and below- ground p a r t s were separated. Old growth (Le., growth occurring in previous years) was removed from below-ground parts.

Biomass samples were dried at 40°C to a constant weight and weighed using an electronic Yettler balance to the nearest 0-1 g or 0.001 g depending on total weight (Le., if less than ' 160 g then weighed to 0.001 g ) to obtain both above- and below-ground d ry weights. This information was also used for nitrogen budget calculations (Chapter 2).

In addition to destructive biomass sarnpling, water depth, number of cattail and buirush shoots, and the presence of other species was recorded for each plot, Also, the number of cattail infloresences w e r e counted for each zone rather than per plot.

4.2.3 ûecom position

Decomposition was assessed using litter bags made of f ibre glass window screening (Davis, 1991) exposed for various periods throughout 1992 to 1994 (see Table 4.2 for experimental design). The window screening was sewn into "litter bags" approximately 10 x 12 c m with monofilament line. Each bag was weighed and fiiled with 4 g (1992) or 1 g (1993/94), dried and chopped leaves (approxirnately 0.5 to 2 c m in length to standardize surface area), to the nearest 0.001 g. Green leaves were used in 1992 and 1993 (after Bayley et al., 1985; Nelson e t al.. 1990a and 1990b). However, weight loss was much higher than many studies reported in the fiterature because of the difference in Litter quality. Standing dead Ieaves were used in 1994 (after Bruquetas d e Zozaya and Neiff, 1991: Davis. 1991) to address this. The bags w e r e tied to stakes and dlowed to fioat for the f i r s t few days until they became water logged and settled to the sediment surface. New litter bags were placed in the trenches each year. One set of replicate bags from each station at each site w e r e removed periodically af ter initial placement, Upon retrieval, the bags w e r e rinsed to remove blue-green algae, various invertebrates, and roots and shoots that had penetrated or covered the bags. Bags were then dried to a constant weight at 40°C and weighed on an electronic Xettler balance to the nearest 0-001 g.

'7

Production data ( biomass/m') were logged [ Le., log( biomass+ 1) ] follow in g convention for biological data containing zero values to produce approxirnate normal distr ibution (Zar. 1984). Decomposition data (as percent of weight lost dur ing exposure period) were arcsin square root transformed following convention fo r percent data (Zar, 1984). -411 statistical analyses were performed on t h e transformed da ta using the SYSTAT statistical package (Wilkinson, 1990). A probability of a c 0.03 was used t o define statistical significance.

Repeated measures analysis of variance (ANOVA) was conducted on t h e constructed wetlands data set, to assess the effects of wastewater type. However, 1992 data were not of repeated rneasures design and were treated as one-way M O V A design. Although the re were three to five replicates (e.g.. quadrats or l i t ter bags) for each zone within each sampling period these individual data points were not used. bu t ra ther t h e means of these data were used. I t is valid t o conduct these statistical analyses over means especially if t he re are rnissing data (Hudson, pers. comm.). Yissing data were especially prevalent fo r t h e decomposition data set in t h e longer exposure periods. Individual t renches were used as replicates of t h e wastewater type. kieasurements were replicated th ree times in each sampling period (i.e.. date) o r zone (i.e.. in. mid. out) . The treatments (Le., wastewater type) were sampled in triplicate and these same replicates (Le.. t renches) were also sampled repeatedly over tirne (Le.. num ber of sampling periods) and space (i.e.. t h r e e zones within each t rench) . Repeated measures design is t h e same as a spli t plot design. with t he treatments applied to the whole plots o r t renches and t h e sampling period and zone applied within the plots o r t renches (Le.. t h e trenches were replicates for tes t ing treatments. but blocks for testing t h e sampling period and zone effects) . No trench effects were assumed.

Post-hoc contras ts were used to determine w hich means were significantly different using Bonferroni ad justed alpha levels ( Le.. critical ievel, 0.05.

divided by t he number of m n t r a s t s tes ted) . Testing a s u b s e t of corn parisons ra ther than al1 possible pairwise corn parisons ( e.g.. Tu key's HSD) often gives more power to the analysis (Day and Quin. 1989: Wilkinson, 1991).

4.3 Results and Discussion

4.3.1 Production

Corn parison Among Treatmen ts

Total production (i.e.. peak above-ground s tanding crop) varied significantly among treatments in 1993 only (Table 4.3). In 1992, no differences among treatments w e r e detected by one-way ANOVA on the data collected from the constructed wetlands ( n = 3; p > 0.05). In 1993, cattail above-ground biomass w a s higher in t h e Control and/or Dyke Drainage t renches than in t h e Pond t r enches (F = 5.48: p = 0.04). Bulrush biomass was higher in the Dyke Drainage and/or Pond t renches than in t h e Control t renches (though not statistically significant). These t rends were most evident in September 1993 (F igu re 4.1). In addition, above-ground biomass for other macrophytes tend to be g rea t e r in t h e Control t renches than in t h e Dyke Drainage and/or Pond t r enches ( though not statistically significant). However. total biomass w a s no t significantly differen t ( F =

2.137: p = 0.197). And again in 1994. above-ground biomass was not significantly different (e.g.. total biomass: F = 0.305: p = 0.59) between treatments (Le., water input type) .

The lack of difference among treatments sugges t s t h a t oil s a n d s water w a s not affecting macrophyte above-ground biomass. although the amount of variability could have masked differences. For example. in 1994 data, a posteriori power analysis indicated tha t between 14 and 51 sarnples would have been required to e n s u r e only 20 % (Le., power of 80 %) chance of

committing a Type I I e r r o r (i.e.. accepting t h e nul1 hypothesis when it should not have been accepted ). Although t renches were s u bsampled. t he re were only th ree t renches per treatment and a total of 9 samples per t rench were collected.

Trends in above-ground biomass were not consistent over al1 s tudy years . Although in 1993 Typha latfloiia biomass was higher in t h e Control t renches than in t he Dyke or Pond t renches, and o the r species biomass aiso tended to be higher in t he Control t r enches (although not statistically significant). i t i s not possible from this s t u d y alone to sugges t the fate of macrophytes under long term exposure to oil s a n d s wastewater. There were d a t a ( though rarely statistically significant) t h a t indicate t h e potential for lower production and diversity in wetlands receiving wastewater.

Potentiaily, production of Typha latifolia and other species could be higher

and Scirpus validus lower in the Control trenches compared with Dyke and Pond trenches (e-g., Figure 4.1). &th biomass and number of taxa of other species tended to be higher (though not statistically signif ican t ) in the Control trenches. However, the response of lower species richness and the same or higher abundance is a typical cornrnunity response to environmentai pollution (e.g., Pearson and Rosenberg. 1978). This trend in diversity was also observed in the Natural Wetland when cornpared with the Reference Wetlands. The Natural Wetland w a s dominated by Typha latifolia and Carex sp., while several more species were present in the Reference Wetlands. This sugges t s that treatment wetlands may end u p w i t h limited diversity. The impact of decreased diversity on treatment effectiveness/sustainability is unknown. Should constructed wetlands Pulfill a reciamation requirement as well as a treatment function, lower rnacrophyte diversity could impact wildlife usage.

Below-ground biomass w a s highly variable and no trends were observed among treatments (Table 4.4).

Variation Bet ween Years

Production between 1992, 1993 and 1994 was significantly different for Control and Dyke trenches. Production of the Control and Dyke trenches tended to be lower in 1994 cornpared with 1992 and 1993 ( F = 17,372; p =

0.001). However, production between 1992 and 1993 w a s not different for Control, Dyke and Pond trenches.

1 1

In 1993, T. latifdia mean biomass was 373 g m-- and 265 g m-' in Control and Dyke Drainage trenches, respectively. In 1994. T. latifolia rnean biomass waç 140 g and 176 g in Control and Dyke Drainage trenches, respectively. This same trend of decreased biomass in 1994 (as compared with previous years) was also observed for S. validus biomass.

Decreased p rod uct ion over time may a f fec t treat men t effectiveness/sustainability and wildlife usage. Production decreased in ail

constructed wetland t renches between 1993 and 1994. This was attributed to muskra t s ( r a the r than application of oil sands wastewater), as muskrats

were observed in al1 trenches. There was physical evidence of muskrat use of cattail and bulrush shoots. Muskrats are known to cut large amounts of emergent vegetation, primarily cattails (Lacki e t al., 1990). and build rnuskrat "houses" and alter the landscape pat terns of wetlands (e.g. . ?litsch and Gosselink, 1993) often referred to as "eat-outs" ( e-g.. Errington e t al., 1963). Muskrats can be a major concern with respect to treatment wetlands because they can reduce dense s tands of emergent rnacrophytes (e-g. , Berg and Kangas, 1989 found they cut about 20 %) to a patchwork of open water and vegetated areas, and can burrow into dykes, which couid lead to system failure (Kadlec and Knight, 1996).

In 1994, muskrats were seen daily. There was evidence of muskrat cuts through the trenches, especially the Control trenches, and the beginnings of muskrat mounds/houses. Muskrats appeared to prefer the Control to the Dyke trenches which were preferred over the old Pond trenches that now

received Dyke water, as almost al1 bulrush and significant number of catîail shoots had been cut from t w o Control and one Dyke trenches. Although muskrats were regularly trapped and removed to o ther locales, muskrats were observed in the s tudy area throughout the 1994 s tudy year and by the end of 1994 there w a s evidence of muskrat cu t s in al1 trenches.

This relativeiy significant impact on the emergent vegetation did not appear to impact treatment effectiveness. However, if this Ievel of muskrat activity continued treatment could become lirnited. Therefore. the potential of muskrat activity in any potential constructed wetlands should b e considered during wetland design to prevent any reduction in treatment effectiveness over t ime .

Cornparison w i t h :Vat ural ketlands and Reference hetlan ds

Above-ground biornass in the constructed wetlands (334 g rn-' to 676 g rn-') was comparable with the biomasç observed in t h e naturai wetlands (398 to

691 g rn-') and reference wetlands (242 to 982 g rd).

Although Scirpus vaiidus was present in the adjacent Natural hetland, the Natural Wetland w a s dominated b y Carexspp. and T w h a latifolia. However. in the constructed wetland trenches, in 1993, T,vpha latifolia tended to predominate in the Control and Dyke trenches, while Scirpus validus tended to predominate in the Pond trenches. As the Yatural Wetland had

been receiving Dyke water since the mid 1980s, it appea r s tha t other

factors rnay impact macrophyte communities in t h e long term.

Water levels in the Natural Wetland increased about 15 cm in 1991 due to construction of a berm at t h e outflow end of t h e Natural ivetland. Since 1991. macrophyte biornass has decreased possibly d u e to the change in water level, making it difficult to determine if Dyke water may deleteriously impact aquatic macrophytes. While the a g e of wetlands can also affect t he production (Noe et al., 1995) and the reference wetlands were likely older than the Natural Wetland. the production of t h e Xatural ketland was higher possi My due to nu t r ien t enrichment from s u bsurface seepage from the adjacent tailings ponds.

Although above-ground biomass decreased over time in t h e constructed wetland trenches. biomass tended to remain higher than in t he Reference Wetlands. Biomass in t h e Reference Wetlands was generaliy lower than in oxbow lakes of Athabasca River. The average peak above-ground biomass

1

in oxbow lakes of Athabasca River ranged from 456 to 848 g m-' in 1981

(Lieffers. 1983). One wetland in t h e 1981 study was t h e same as used in th i s s tudy in 1991 (Le. t h e Shipyard Lake Reference Cv'etlands). The average above-ground biomass in 1981 was 456 g rn-' (Lieffers. 1983)

1

compared with 373 g m-' in 1991 ( this study).

4.3.2 Deam position

Corn parison among trea tmen ts

Decomposition (as measured by % weight loss) occurred in al1 treatments,

but tended not to Vary significantly among t reatments, except in 1993 (Table 4.6)- In 1993. decomposition was significantly di f ferent between treatments ( F = 22.2: p = 0.002). and dl t reatments differed significantly

from one another (F = 37.2: p = 0.001 as determined by post-hoc testing - linear contrast). In 1992, decomposition was lower in Pond trenches than

in either t h e Dyke or Control trenches, but the difference was not significant among t reatments (F = 1.14; p = 0.38). In 1994. decomposition w a s not significantly different among treatments for e i ther T. lafiT01ia ( F

= 0.46: p = 0.55) o r Scirpus validus ( F = 1.41: p = 0.32).

Generaily, decornposition tended to decrease from Control to Dyke to Pond trenches (Table 4.5). Although these trends were only significant in 1993

(Figure 4.2), they suggest that leaching and/or bacterial decomposition of leaf litter w a s retarded under the heavier contaminant loads that occurred in 1993 (lie., 6 L/min inflow in 1993 cornpared with 2 L/min in 1992 and 1994). This s tudy could not determine if t he reduced decomposition was d u e to cornpetition for resources (e.g-, oxygen, phosphate) or the toxic effects of the contaminants.

In study years where no significant differences were observed among treatments a posteriori power analysis w a s conducted. Ignorance of a study's power can have serious mnsequences by misleading environmental managers preparing mine closure reclamation plans (Peterman, 1990). Power

was low in 1992 and for longer exposure periods (Le., > SI days) in 1994. Overall power was acceptable for shor ter exposure periods, but not for longer. Between 14 and 51 replimtes o r more would have been needed io improve power to acceptable levels (i-e.. 80 % ) for longer exposure periods,

Li t ter Qualits'

Several factors can affect decomposition: ciimate ( Le.. temperature, moisture). litter type, oxygen. nutrients and flow. khile decom position is

slower under cooler temperatures (Nelson et al., 1990a). temperature effects were not observed in this s tudy and within year differences in temperatures were not tested. Litter quality also affects decorn position rates (Selson et al., 1990a: Szumigalski and Bayley. 1996). Generally. t he higher the nutrient content the €aster the rate of decornposition. Szumigalski and Bayley (1996) found that percent weight loss was correlated with N content and C:S ratios (Le., t he greater S t h e greater the weight loss af te r 1 and 2 years) .

In 1993, decomposition occurred in al1 treatments a t a higher rate than in 1992, while in 1994 decomposition occurred in al1 treatments a t lower rates than in 1993. L i t t e r type w a s different among s tudy years: green in 1992 and 1993, and standing dead in 1994. Flow was different among s tudy years: 2 L/rnin in 1992 and 1994, and 6 L/min in 1993. Either of these factors could be responsible for the differences observed.

As there w e r e no differences in clirnate or li t ter type between 1992 and 1993, it is possible that the change in decomposition was d u e to changes in flow or the available nutrients. The increased Flow rate may have increased dissolved oxygen. The increased Flow in 1993 increased the

contaminant loading which may have increased the demand for oxygen and nutrients, but also m a y have increased the dissolved oxygen content. Since t h e decomposition rates increased between 1992 and 1993. it is also possible tha t t h e microbial populations were not well established in 1992 compared with 1993 populations, which would have increased decomposition in 1993.

While flow rates affect decomposition, t h e resuIt varies depending on the wetland examined (clitsch and Gosselink, EX?). Decorn position may occur more rapidly unde r periodic flooding (Brinson e t al.. 1981) or in the mainstrearn of a r i v e r r a t h e r than on the w e t swamp floor (Brinson. 1977). Lnder permanently s u bmerged conditions t h e r e is be t te r access to aquatic detritivores, a more cons tan t physical environment For t h e microbiota. a greater availability of d issolved nutrients, and a bet ter environment for leaching (Odum a n d Heywood, 1978). However, if water is anaerobic biological activity can be significantly reduced (Tupacz and Day. 1990).

Decomposition of Typha latifolia was 52 % and 47 % in 1993 for the Control and Dyke trenches. respectively; compared with 19 % and 18 % in 1994. These t rends were not tested statistically due to t h e differences in experimental design. The differences in decorn position between 1993 and 1994 was at t r ibuted to t h e change in Iitter quality. Standing dead leaves were used in Iitter bags in 1994 ra ther than green leaves as in 1993 and 1992. Litter qual i ty of the standing dead leaves is lower ( e . . the proportion of the h a r d e r to decompose lignin fraction is h igher ) than in green leaves. Hig her decorn posi tion in green leaves corn pared wi th standing dead leaves has been observed elsewhere (Sclson et al.. 1990a).

The simple exponential rnodel (Q/$ = dt) described by Olson (1963) w a s used to estimate rates of decay. This rnodel is widely used where the relative decay rate remains constant through time assuming easily degraded and recalcitrant materials interact to yield a constant decay rate (Kelson. 1990a). Decay rates a n d haif lives are given in Table 4-5 and can be compared with l i t e ra ture values presented in Table 4.6. Half lives ranged from 39 to 347 d a y s (calculated from decomposition in t h e f i r s t 42 days), and were within the ranged of 29 to 610 days reported in t h e titerature, Nelson (1990a) estimated decay rates for green ieaves at about twice the ra te as for senesced leaves. In this s tudy the difference was about three times for Typha lat1Toiia Decay ra tes calcuiated from data given by Thormann and Bayiey (1997) for green Typha lat30Iia leaves in a rnarsh

north of Edmonton were 0.0154 and 0.0028 a f te r 30 and 363 days. respectively. Decay rates af ter t h e shor te r period were comparable with 1993 rates in this study sugges t ing that the difference between years w a s due to litter quafity.

Species Varia tion

Differences between decorn position rates for Typha latifoiia and Scirpus validus may also be a t t r ibu ted to differences in litter quality. Different decomposition rates fo r Typha spp. and Scirpus spp. have b e n observed in other studies (e.g., Kulshreshtha and Gopal (1982) found tha t Scirpus sp. decay rates were almost twice as fast a s for Typha sp.) . Although the differences between t h e s e species were lower. they were kikely a t t r ibutable to differences in l i t ter quality. Nitrogen fias been sugges ted to enhance decomposition rates (Puriveth. 1980: Thormann and Bayley, 1997). Scirpus validus t issues tended to have higher nitrogen con tent than Typha 1atifoIia (Chapter S ) , suggest ing that nitrogen content affected decomposition ra tes

in this s t udy also.

Cornparison with .La tural Cu'etland and Reference hetlands

The trend of higher decomposition in the Control compared with D x k e and Pond water t renches was comparable with the t rend of higher decomposition in the Reference Wetlands than in the Yatural ketland (Figure 4.3). Decomposition w a s significantly lower in t h e Xatural Netland than the Reference Wetland in 1993 ( T = -2.5; p = 0.022). though significant

t r ends were not observed in other years.

Decomposition w a s also f a s t e r in the Reference Wetlands and Natural Netland compared with t h e constructed wetland trenches. This difference was likely due to winter temperatures. Water Flow to the constructed wetland trenches did not continue throughout the winter and t h e t renches were allowed to freeze over t he winter. However. t h e Natural Cvetland received continuous inpu t s and did not freeze and it is presumed also that the Reference Wetlands also did not freeze to t he sediments.

4.4 Conclusions

Although the potentiai impacts of oil sands wastewater (Le.. Dyke and Pond water) on aquatic macrophytes are still not well understood. t he re is

potential that both macrophyte production and decomposition may be negatively impacted by oil sands wastewater. As significant differences were only observed in 1993 (unde r higher loads), it is possible impacts can be controIled by controlling loading. Additionaliy, natural impacts such as t h e effect of muskrats o r beavers may significantly affect the success of treatment constructed wetlands. A plan to address them should be included in t h e design, implementation and operation of treatment wetlands.

The most significant observation is that macrophytes continued to grow and appeared to "thrive" in t h e oil s ands wastewaters af ter three s tudy years of appiication. Given t h e high natural variability in production in wetlands, the effects of climate, muskrat ôctivity and wastewater loading cannot be separated. Macrophyte production tended to be lower in t renches receiving oil s ands wastewater, though not significantly. However. t h e samplîng design may have been inadequate due to relatively high variation a n d a low number of replicates as suggested by t h e a posteriori power analysis. The lack of significant differences suggest no impact: however, t h e power of t h e s t u d y w a s low so the potential for accepting t h e nul1 hypothesis when it was false w a s high. Larger-scale,

longer-term demonstration wetlands are required before im plemen ting treatment wetlands as par t of mitigation design/mine closure.

Leaf l i t ter provides sur face area and nutr ients for bacteria, and therefore

can indirectly affect contaminant rernoval. I n addition, decomposition plays an important role in overail wetland function by recyciing nutrients- Although t rends in decomposition were not sbt is t ical ly significant in ail s t u d y years , t he re was some indication tha t oil sands wastewater could reduce l i t ter decomposition. A posteriori power analysis indicated tha t t he re were insufficient replication at longer exposure periods. while at shor te r exposure periods i t was acceptable (Le., power of 80 %).

Differences in decomposition under sho r t e r exposure periods were only detected under t h e higher loading in 1993. However, the cause of t he observed decreased decomposition rates could have been t he higher contaminant loads o r low dissolved oxygen o r some combiriation.

The impact of reduced macrophyte production and/or decomposition rates on treatment efficiency and sustainability i s unknown. In 1994, when t h e

apparen t production was substantially reduced by muskrats, wastewater treatment efficiency was not affected (Chapter 2). While no clear t rend of contaminant impact on macrophyte production and/or decomposition was

observed , i t is important to note that loss of wetland function could limit the use of wetlands as a reclamation option and perhaps atso as a treatment option for the oil sands industry. This study suggests that a longer term larger, pilot-scale project is warranted to assess the efficiency of wetland treatment and the sustainability of wetland function.

Table 4.1 Experimentai design for production measuremen ts.

- -

Design Feature

Sites

Zones per Site

Quadrats per Zone - Quadrat Size (m2)

random random selec ted h Quadrat Location Determination

Sam ple Periods July August S e ~ t e m ber

Parameters (frequency sampled ) A bove-ground Biomass Shoot Density Shoot Lengths Shoot Widths Diversity Typha IaWoIia Inflorescences Below-ground Biomass hater D ~ D tn

LX Cons t ruc t cd kc t 1 ands Trcnchcs Rh Rc f c rcncc hc t lands ?'

Natural kct lands

nnt mcits i i rrd mciisii rrd

a 1992 samplc pcriods: abovc-~roiind biomass <in .lui' 30 :IU~US t TO- L3

b 1943 samplc pcriods: Junc 26-JuLy 3: August 1 - 7 : S e p T9 werc snmplcd in Augus t (2nd l sumcy

C: 1994 s m p L c pcriods: Juiy 6-7: .lu~ust 3 - 4 : Scptcmbcr d sec Appcndiccs X e spi i t t rcnchcs wcrc mnni t o r d onlrtl in Scptcmbrrr 1!194

(6 rcplicatc and 6 spiitl

- 3 1 : n L nc r pn ramr : i-rs

tcmhvr 1 - 9 : only T7. TH and

i - ! 3

for a i o ~ a i of' 12 Lrcnchcs

r 1 for abovc-uround biomnss and 3 for othcr pnrametcrs mcasurcii R h

I for ahovc-%round biomass and 5 for othcr parmctcrs mcasiirt?d sclectcd boscd on sparsc vcgctation so largc zoncs sclcctcd for shoot dcnsity and morphology?: and better establishcd arco ( @ Stacion 8 ) sclccted for biomass to prcvent dccimation of a particular arca

Table 4.2 Experimental design for litter decomposition-

Sites

Zones per Site 3 3

Litter Bags p e r Zone

Initial Placement August 15 July 24 June 26

Exposure 1 sd 3e Per iod s

S pecies

a b

Grave1 Fit Reference KetIand see Appendi.. A originally placed three bags per zone, but two bags per zone were used for rnicrobiological tests

C 4 1 da)- exposure period 9 21 day, 42 day. 63 d a s . 290 day, and 438 day exposure periods e 21 day. 42 day, and 84 day exposure periods

Type

Total Nurnber of Litter Bags

Typha latifolia

Green

27b L

Typha la tifolia Typha latifolia Scirp us valid us

Green

825

Senesced

720

Table 4.3 Above-ground macrophyte biomass (meari 2 sLari(lai-tl c h ia t ic $ 1 1 ;

in the constructed wetlands, 1992 - 1993.

Septem ber 1 994

1

Dyke Seepagc - , GL-,- - 1 6 2

Pond I A jj; = 52 Naturd b'et lands 534 = 1 Ji

Retérence D'etlands 37; = lS3

Dyke Secpare 285 = l l t j

Pond i .A 221 = 1 % Natural Wetlands 130 = 189 Rrference L i et lands -3.1 = 6.3

Control 37; = 27L,

Table 4.4 Below-ground macrophyte biomass (mean + standard deviation) in the constructed wetlands, 1993.

Treatmen t Dry Weight (g m-') Ratio Below :.A bove

Typha Scirpus Total Total la t ifolia valid us Below - A bove-

ground ground

Control 2841375 69722413 98 1 5 18 1.9

Dyke Drainage 3562423 581268 414 423 1.0

Pond 1A 30614 11 5002945 806 432 1.9

Nat ural 2062179 020 206 176 1.2

Reference 841161 12139 96 1 05 0.9

Decomposition ( % reduction from initial mass, mean 2 SD) of Table 4.5 le& litter over various exposure per iods from 1992 - 1994.

l'reahen t Exposure Period (days 1

. ., -.

Cont r d 1928 0.005 139

, @ke 19CN 0.005 139

Pond 9i9 0.002 347 l

1993 [green mpba Iatifolial

1994 (standing dead opha lati folial

a actualiu onl. 4 i dq esposure period in 1992 5 detewined using data from 42 Say exposure

1994 (standing dead Srirpus raiidus)

ihere : & = i ~ i t i a l iass Qt = iass a i t h e t t = t h e , days k = decal coefficient. da!-'

(Olsoa, 1963) c calcuIated using decaj- coefficient detmined fro 12 day exposure - nodata

48.3i6.4

48.8t3.6

39.823.2

Controi

ûyke

Pond

30

46

69

61.8126.6

31.511. 1

38.193.3

0.005

U.005

23.3t8.9

?3.5?8.8

Cantrol

. hie

44.714.0

9 . f 3 . 2

28.913.3

139

139

U.1i07

il. $08

32.8213.9

37.8r8.2

Cantrol

Dvke

52.2t9. I

47.3t4.2

33.624.9

6 6 . 2 2 2 . 6

60t27.7

14.523. 4

12.8!:.6

99

4:

fl.418

0.013

0.010

18.9i6.2

17.7t7. i

1:.42.4

20 .8 t l l . J

24.311

2 0 . 6 5 . ;

Table 4.6 Decay coefficients and half Iives for litter under various water quality conditions.

Species Water Study Decay Half Li fe Du ration C o e f ficiev t ( days Reference

(dl (k. day-')

T Y P ~ ~ Natural 30-60 0.0033- 197-393 Hcrshkowi r z . 13Ht;

SPP* 0.00 18

45 Thormann and

248 Bayiry. 1997

138 0.0240 29 Sr i son. 1990a

(green (green) 0.0110 63

( senesced ) ( senesced )

300 0.0104 67 h r b s t c r and Simmons . 197H

haste- 30-60 0.0074- 93- 186 Kadlcc. R . H . .

water 0.0037 iYH6

Bath n 0 t 0.00 17 400k250 Iiadlt-r R . H . . i 9 8 6

given and Ri11 shrcshtha and Gopai . 19Hf

Scjrpus Satura1 not 0.0027 260+80 hulshrcshtha and

SPP. given (;opol . 1982

Cares Satura1 30-60 0.00 11 610+320 Kadlcc. R.H.

0.0002- 289-347 Thormar! and

0.0024 Ba?- 1 cy . ! 99 7

363 0.0008- 433-866 3zumigtlski and

0.0016 Bayiey. 1996

(senescent )

assumes decomposition rate: Q/Qo = e -kt

Figure 4.1 Above-ground macrophyte biornass (mean i SD) in t he c o n s t r u c t e d wetlands, September 1993.

99

1 993 o p h u lutif uiiu

Green

Kgut-e 4.2 Yass rernaining ( d r y ~ e i g h t as % of' initial mass: mean + 5k. for Q7pha latifdia green leaves in 1993 in t h e ~ o n s t r u c t l - ~ t i wetland t renches .

Figure 4.3 Yass remaining ( d r y weight as 76 of initiai m a s s nwbii i. S E l for Typha latifolia green leaves in 1993 in the Katural Wetland and Reference Wetlands.

4.5 Literature Cited

Bayfey, S.E., Zoltek, J. Jr., Hermann, A.J., Dolan, T.J. and 1,- Tortora. 1985. Experimental manipulation of nutrients and water in a freshwater marsh: effects on biomass. decomposition a n d nu t r ient accumulation. Limno1og.v and Oceanography 30(3):500-312.

Berg, K.M. and P.C. Kangas. 1989. Effects of muskrat mounds on decomposition in a wetland ecosystem. pp. 145-151. In: R.R. Sharitz and J.k. Gibbons (editors). Freshwater wetlands and wildlife. Oak Ridge, TN: L'S Department of Energy. COSF - 8603101.

Boyd. C.E. 1970. Losses of minera1 nutr ients during decomposition of Typha la tifoiia. Arch. Hydrobiol. 66: 5 1 1 -5 17.

Brinson. Y.Y. 1977. Decomposition and nutr ient exchange of Litter in a n alluvial swamp forest. Ecology 61:827-835.

Brinson, K Y . , -4.E. Lugo and S. Brown. 1981. Primary productivity, decomposition and consumer activity in freshwater wetlands. Annu. Rev. Ecol. Systematics 12:123-161.

Brui, H. 1990. Gas exchange through the soil-atmosphere interphase and through dead culms of Phragmites australis in a constructed reed bed receiving dornestic sewage. hafer Research 24(2):239-266.

Bruquetas d e Zozaya, I .Y. and J.J. Neiff. 1991. Decomposition and colonization by invertebrates of T,rrpha Iatifolia L . l i t ter in Chaco cattail swamp (Argentha!. Aquatic Botan). 30: 18% 193.

Chamie. J.P. and C.J. Richardson. 1978. Decomposition in northern wetlands. 1 n: Fresh water Wtlands; Ecological Processes and .Vanagement Potentiai. R.E. Good. D.G. h'higham. and R.L. Simpson. eds. New York: ilcademic Press. pp. 115-130. [cited in Liitsch and Gosselin k , 1993 3

Cooper, P.F. and J.A. Hobson. 1989. Chapter 11. Sewage treatment bb* reed bed systems: the present situation in the Cnited Kingdom. In: Harnmer . D A (ed. ) Constructed wetlands for wastewafer trea tmen t: municipal. industrial, agriculturai. Proceedings of the First International Conference Constructed CVetfands for Ciastewater treatrnent, Chattanooga, TN. Chelsea: L e w i s Publistiers. pp. 153472.

Davis, S.Y. 1991. Growth, decomposition, and nutr ient retention of Cladium jamaicense Crantz and Typha domingensis Pers. in t h e Fiorida Everglades. Aquatic Botany 40203-224.

Day, R.K. and G.P. Quinn. 1989. Cornparisons of treatments after a n analysis of variance in ecology. Ecd. Monogr. 59:433-463.

Desjardins, R.M.. P.L. Seyfried. and G. Palmateer. 1986. Heavy metal effects on marsh treatment systems. kater Poilution Research Journal of Canada 21524-535.

Dunbabin, J.S. and K.H. Bowmer. 1992. Potential use of constructed wetlands for treatment of industrial wastewaters containing metals. The Science of the Total En vironmen t 1 1 1: 15 1 - 168.

Errington, PL.. R.Siglin and R. Clark. 1963. The decline of a muskrat population. J. WiMl. Manage. 27:l-8.

Faulkner. S.P. and C.J. Richardson. 1989. Physical and chernical characterist ics of freshwater wetland soils. (Ch. 4). In: Construcfed wetlands for waste wa fer trea tmen t: Municipal, In dustrial and Agricultural. D.A. Hamrner, Ed. Chelsea, Yichigan: Lewis Publishers. Inc, pp. 41-72.

Gersberg. R . 3 . . B.V. Elkins and C.R. Coldrnan. 1984. Cse of artificial wetlands to remove nitrogen from wastewater. Journal kater Pollution Control Federation. 56: 152- 156.

Gersberg, RA., B.V. Elkins, S.R. Lyon and C.R. Goldman. 1986. Role of aquatic plants in wastewater treatment by artificial wetlands. hater Research 20:363-367.

Gersberg. R.Y.. R. Brenner, S.R. Lyon and B.V. Elkins. 1987. Survival of bacteria and viruses in municipal wastewaters applied to artif icial wetlands. pp. 237-245. In: KR. Reddy and L.H. Smith (editors). -4 qua tic plan fs for water trea tment and resource recovery. Orlando: clagnolia Pu blishing.

Gulley. J.R. and D.J. Klym. 1992. ketland treatrnent of oil sands operation wastewaters. In: R.. SinghaI. A.K. Mehrotra. K. Fytas and J.-L. Collins (eds. ) Environmen W Issues and h a s t e .Vanagement in Energy and Minerals Production. Proceedings: Second International Conference on Environmen ta1 Issues and Yanagement of kas t e in Energy and ."lineral Production, Calgary, Alberta, September 1-4. 1992. A.A. Balkema, Rotterdam, pp. 1431-1438.

Hamilton. S.H. and P.C. Xix. 1991. Wetlands a s a treatrnent system for d y k e drainage water: i ï terature review and feasibility s tudy. Prepared for Suncor Inc.. Oil Sands Croup by EVS Consultants Ltd. 45 pp. + appendices.

Hers h kowitz, J. 1986. Listowel artifiuaI marsh projeet- Report prepared fo r the Ontario Ministry of the Environment. pp. 11 1-116. [cited in Kad lec, 1989 j

Hudson, Gary. pers. comm. Technical representative of SYSTAT. Telephone conversation discussing how to handle missing data in repeated rneasures ANOVA and how t o assess w h ich t rea t men ts. zones o r periods were different. May 19, 1995.

Kadlec, R.H. 1986. ketland Ctilization f o r management of community wastewater: 1985 Operations Summary, Houghton Lake . Yichigan. Report to Michigan DNR. pp. 51-78. [cited in Kadlec, 19891

Kadfec, R.H. 1989. Decomposition in wastewater wetlands (Ch. 37a). In: Constructed wetlands for wastewater trea tmenf: :Vunicipal. Industrial and Agriculfural. D.A. Hammer, Ed. Chelsea, Michigan: L e w i s Publishers. Inc. pp. 459-468.

Kadlec, R.H. and R.L. Knight. 1996. Treafment ketlands. Boca Raton: Lewis Publishers. 893 pp.

Kadlec, R.H and D.L. Tilton, 1979. The u s e of freshwater wetlands as a tert iary wastewater treatment alternative. CRC Critical Reviews in Environmental Control 9:185-212.

Kulshreshtha, Y. and B. Gopal. 1982. Decornposition of f reshwater wetland vegetation. II. Aboveground organs of emergent macrop h ytes. In: Ketiands: Ecology and Management. B. Copal. R.E. Turner , R.C. Netzei, and D.F. Whigham, Eds. Paris: UNESCO. pp. 279-292.

Lacki. M.J.. KT. Peneston, K.B. Adams, F.D. Vogt and J.C. Houppert. 1990. Summer foraging pa t te rns and diet selection of muskrats inhabiting a fen wetland. Canadian Journal of Zaoiog~~ 68:1163-1167.

Lieffers. V.J. 1983. C r o w t h of Typha latifolia in boreal forest habitats, as measured by dou bie sarnpling. Aquatic Botany 15:335-348.

McNaughton. S.J., T.C. Folsom, T. Lee, F. Park, C. P r i e . D. Roeder. J. Schmitz and C. Stockwell. 1974, Heavy metal tolerance in Typha latifolia without the evolution tolerant races. Ecolog~- 53:1163-1165.

Mitsch, W.J . and J.G. Gosselink. 1993. Metlands (2nd ed.). Van Nostrand Reinhold. New York. 722 pp.

Yorea, S., R. Olsen and T. Wilderman. 1990. Passive treatment technology cleans u p Colorado mining waste. Ka ter Environmen f and Techoology Decern ber: 6-9.

Yoss, E.H. 1983. FIora of Alberta. 2nd Edition revised by J.C. Packer. Toronto. ON: University o f Toronto Press. 687 pp.

Nelson, LW., J.A. Kadlec and H.R. Yurkin. 1990a. Seasonai cornparisons of weight loss for t w o types of Typha glauca Godr. leaf litter. Aquatic Botany 37:299-314.

Nelson. J.K., J.A, Kadlec and H.R. clurkin. 1990b. Responses by macroinvertebrates to cattail l i t ter quali ty and timing of l i t ter s u bmergence in a northern prairie marsh. Metlands 10( 1 ):47-60.

Nichols, D.S. 1983. Capacity of natural wetlands to remove nutr ients from wastewater. Water Polluticn Control Federation Journal 55:493-305.

Niering, W.A. 1989. Wetlands. New York, SY: Alfred A. Knopf, Inc. 638 PP-

Xix. P.G., S.H. Hamilton, L. Bendell-Young, C.P. Cunter, F.S. Bishay and Y.D. Paine. 1994. Constructed wetlands for t h e treatment of oil s ands wastewater: Technical Report #3, Prepared for Suncor Inc., Oil Sands Group by EVS ConsuItants, So r th Vancouver, B.C. 223 pp. + appendices.

Noe, C.B., R.B. Atkinson and J. Cairns, Jr. 1995- Primary productivity measured as peak aboveground biomass of emergent macrop hytes in accidental wetlands on surface mined areas- Reclamation .'Vewslet ter Summer 199515-17.

Olson, J.S. 1963, Energy s torage a n d t h e balance of producers and decorn posers in ecological systems. Ernl~g.,~ 443322-331.

Peterrnan, R.Y. 1990. Statistical power analysis can improve fisheries research and management. Canadian Journal of Fisheries and -4quatic Sciences 47(1):2-13.

PuIlin, B.P. and D.A. Hammer . 1991. Aquatic plants improve wastewater treatrnent, kater Environment and Techno1og.r Yarch: 36-40.

Puriveth, P. 1980. Decomposition of emergent macrophytes in a h isconsin marsh. Hydrobiol. 72:231-242.

Rogers, K.H., P.F. Breen and A.J. Chick. 1991. Sitrogen removal in experimen ta1 wetland treatment systems: etridence for t h e role of aq uatic plants. Research Journal kater Pollution Con trol Federation 63(?):934-941.

Szumigalski, A.R. and S.E. Bayley. 1996. Decomposition along a bog to rich fen gradient in central Alberta, Canada. Can. JI Bot. 74573-581.

Tanner, C.C., J.S. Clayton and M.P. Ubdell. 1995. E f f e c t of loading r a t e and planting on treatment of dairy farm wastewaters in constructed wetlands. 1. Removal of oxygen demand, suspended solids and fa-: coliforms. Water Research 29(1):17-26.

Thormann, M.N. and S.E. Bayley. 1997. Decomposition along a moderate- r ich fen-marsh peatland gradient in Bored Al berta, Canada. Wetlands 17 (1):123-137.

Tupacz, EX. and F.P. Day. 1990. Decomposition of roots in a seasonally flooded swamp ecosystem. Aquatic Botany 37: 199-214. (cited in Kad Iec and Knight 1996).

Van Oostrom, A.J. and R.N. Cooper. 1990. Yeat processing effluent treatment in surface-flow and grave1 bed constructed wastewater wetlands. In: P.F. Cooper and B.C. Findlater (editors). Construcfed wetlands in water pollution con trol. Proceedings Internat ional Conference on t h e U s e of Constructed Wetlands in Water Pollution Control, Cambridge, C.K., Septem ber 1990. Pergamon Press . Oxfci-d. pp. 321-332.

Watson, J.T., S.C. Reed, R.H. Kadlec, R.L. Knight and A.E. khitehouse. 1989. Performance expectations and loading rates for constructed wetlands. In: Hammer, D.A. (ed.) Cmstructed wetlands for wastewafer - treatmen t: municipal, industrial, agricultural. Proceed ing s of the First International Conference Constructed Wetlands for Wastewater treatment, Chattanooga, TN. Chelsea: Lewis Publishers. pp. 319-351.

Webster, J.R. and G.M. Simmons. 1978. Leaf breakdown and inver tebrate colonization on a reservoir bottom. Verin. Limnol. 20: 1587- 1596.

keinnmann, F., M Boulé, K. Brunner, J. Yalek and V. Yoshino. 1984. ketland plants of the Pacific North west. Seattle, KA: U S Army Corps of Engineers. 83 pp.

Cteller. Y . . 1990. Waterfowl management techniques for wetland enhancement. restoration and creation useful in mitigation procedures. pp. 517-528. In: J.A. Kusler and X E . Kentula (editors). Wetland creation and restoration, the status of the science. Island Press: Cu'ashington, DC. 591 pp.

Wilkinson, L. 1990. SYS7XT: The system for sbtisfics. Evanston, IL: SYSTAT Inc.

Zar, J .H. 1984. Biostatistical Analysis. 2nd Edition. Pren tice-Hall, Englewood Cliffs, SJ. 718 pp.

5.0 CONCLUSIONS

5.1 Scope of the probIern

Suncor inc., Oil Sands Group (Suncor) is t h e f i rs t of two mmpanies mining oil s ands in north eas te rn Alberta, 23 km north of Fort YcYurray. The mining and s u bsequent conversion of oil s a n d s to synthet ic c r u d e oil began in 1967. There a r e several environmental concerns associated with the production of synthet ic c r u d e oil. This s t u d y focused on the problem of t h e accumulating fine tailings.

In t h e p r e s s of extracting bitumen (Le.. the Clark Hot hater Extraction Process) from t h e oi1 sands , a waste product called f ine tailings is produced. Fine tailings are a n aqueous suspension of sand. silt , clay and residuaI bitumen and naphtha with a pH between 8 and 9 (FTFC. 1995a). Fine tailings have been s tored in large clay iined tailings ponds to date, which must be reclaimed to a viable land sur face o r water body free of long term maintenance (Nix and Martin. 1992).

There a r e two reclamation scenarios: d r y landscape and w e t landscape. There is a potential for water release €rom both d r y and wet landscape scenarios. Although the water auali ty of t hese released waters is expected to be good (FTFC, 1995b), t h e contaminant ioad is unknown. To a d d r e s s this potential water release e i ther a collection systern must be maintained or a treatment system created. Wetlands could act as a polishing treatment system for water leaving t h e reclaimed site before entering the receiving environment. The potential for th is w a s investigated in this s tudy .

Organic (e.g., naphthenic acids) and inorganic (e-g.. ammonia) contaminants are of mncern in both waters tested (Dyke and Pond water). These contaminants (Le.. naphthenic acid and ammonia) a r e of concern as they a r e toxic to aquatic organisms. The treatment effectiveness of wetlands to retain/remove ammonia and nap hthenic acids (as measured by total extractable hydrocarbons o r TEH) w a s assessed.

5.2 The potentiai for the use of wetlands to treat oil sands wastewater

The th ree previous Chapters addressed t h e potential for the use of wetland t o treat oil sands wastewater by assessing treatment efficiency (Chapters

2 and 3) and wetland viability (Chapter 4). The potential of wetlands treatment was reviewed assessed in the proceeding chapters by addressing five issues below.

5.2-1 Do wetlands treat oii sands wastewater?

The contaminant concentrations were lower in the outflow water than inflowing water for Dyke and Pond water trenches. Both ammonia and TEH concentrations decreased from inf low to ouff low .

Dyke inflow ammonia concentrations were in the range of 10 to 15 mg L". 1

and Pond inflow concentrations were in the range of 10 to 12 m g L? Average Dyke outflow concentrations w e r e in t h e range of 0.15 to 12 mg L", and Pond outfîow concentrations were in t h e range of 0.25 to 10 m g L' '. Cnder comparable loading rates theçe outf low concentrations were comparable w i t h the range of 0.2 mg L" tu 48 m g L" reported in the literature (Hamrner and Knight. 1994).

Dyke inflow TEH concentrations were in the range of 13 to 38 mg L-', and Pond inflow concentrations were in t h e range of 12 to 46 mg L Average Dyke outflow concentrations were in the range of 4 to 28 mg L-', and Pond outflow concentrations were in the range of 8 to 26 mg L-'. These oufflox concentrations were higher than t h e range of 0.1 mg L-' to 2.0 mg L-I

reported in the literature for oil and grease (Table 3.3).

5.2.2 How well do weüands t reat this wastewater?

-4lthough ammonia removal was substantial (ranging from 32 - 99 % ) there was often still an order of magnitude difference between the Controi trench outflows conipared with either the Dyke or Pond trench outflows. In 1992 and 1994, removal rates were in the range of 44 % to 99 %; however. in 1993 removal rates were much lower (20 - 40 %) presumably because of the increased amrnonia loading. A t loading rates of 190 mg m" d-l removal rates averaged > 90 %, while for higher loading rates of 270 mg m - 2 d - l

removal ra tes averaged 75 %. Even iower removal efficiency (32 to 42 %)

was observed for higher loading rates (605 to 785 m g d-'1, although similar loads were removed.

Arnrnonia was effect ively removed by cons tructed wetlands receiving oil sands wastewater when compared with a ther treatment wetlands, despite

receiving a high load of hydrocarbons and being situated at a northern latitude. O bserved data were comparable w i th predicted data using three ammonia removal models ( H a m m e r and Knight. 1994; Reed et al.. 1995; KPCF,

1990). Although observed values may have been statistically signif icantly different from the predicted v d u e s (12 of 15 data sets). amrnonia removal w a s considered effective as the observed ammonia removal was often greater than the predicted removal (8 of 12 data sets).

Average TEH removal rates (Le.. mass reductions) ranged from 19 % to 76

% for Dyke water and 54 % to 69 X for Pond water during t h e three year s tudy. These rates were lower than rates for oil and grease reported in the Literature (Table 3.3). The difference w a s probably due to the amount of contaminant loading. Contaminant loading was on average higher in this s tudy ranging from 146 to 1143 mg d-l compared with 32 to 168 mg rn-' d-' for oil and grease removal in other studieç.

5.2-3 Where do the mnt;iminants go?

in 1994, the oniy year with a complete budget for ammonia nitrogen, the difference between infiow and outflok ( Le. the amount of ammonia "retained" by the wetlands) w a s 4.8 k g trench- ' season-! (206 mg rn-' d - i ) , while the sum of the removal processes and change in nitrogen storage w e r e only 2.35 k g trench-' season-' (or 101 mg rn-? d-i) . Plant storage did

not contribute s u bstantially to ammonia removal. Removal processes such as volatilization and denitrification, when measured in 1994. did n o t account for al1 the ammonia retained.

Ammonia removai to the sedirnents were 49 % (2.3 k g t r e n c k i season" or 101 mg m-' d-l) of the ammonia nitrogen retained (or perhaps more if

immobilization occurred ). while plant uptake accounted for approximately 2 % (0.10 k g trench" season-': 4.5 mg m -2 d- l , . Nitrogen removal by plant uptake is not considered to b e a main removal processes (van Oostrom and Russell, 1994; Richardson and Nichols, 1985; Klopateck, 1975 and 1978 cited in Richardson and Nichols, 1985: Gers berg, 1986).

Volatilization and denitrification accounted for a p proximately 0.1 % (0.004 k g trench" season-' o r 0.19 m g rn'2 d-l) and 4 % (0.19 kg trench-' season-* or 8.16 mg me2 d-l) of the ammonia retained, respectively. Volatilization of ammonia may account for losses u p to 100 mg m" d- ' at pH 8 ( based on an empirical equation developed for predicting arnmonia volatilization from

flooded rice fields by Freney et al., 1985): however, using this study's data in the mode1 3.7 mg me' d-l of ammonia muid have been voiatilized. Although denitrification is often assumed to be the primary remoc-al process (Nichols, 1983). in this study it may only represent a t most 50 % of the observed removal. However, denitrification (as measured ) could not be verified experimentally to be the predominant pathway possiblÿ due to the limitations of the technique rather than the lack of occurrence of denitrification (Seitzinger et al., 1993: Daisgaard and Bak, 1992).

TEH rernoval to the sediments accounted for less than 30 % of the TEH load retained by the wetlands in this study. ûther removal rnechanisms include volatilization, plant uptake and microbial mineralization. Volatilization was not expected t o be significant and paraiiei studies did not indicate plant u ptake w a s significant and the removal via mineralization was unclear. Further method development is required to ensure processes and storage are ref lected accuratel y.

Although some questions remain about the fate of both ammonia and TEH, removal w a s substantial. Further investigation to assess the role of constructed wetlands in the oil sands industry with respect to treatment and reclamation is warranted.

5.2.4 What factors affect weüand treatment efficiency of oii sands wastewater and can these factors be optimized?

Ammonia loading w a s observed to influence ammonia removal rates. Year 1 (1992) had a relatively low loading rate which resulted in a high treatment efficiency {Le.. percent removal). Year 2 (1993) had a very high loading ra te which resulted in a low treatment efficiency. Year 3 (1994) had an intermediate loading ra te which resulted in a n intermediate treatment efficiency. The same trend was observed for TEH removal. However, it may not just have b e n the ammonia and TEH loading rate responsible for the difference. Generally, with higher ammonia Ioading rates corne higher TEH loading or visa versa. and for both lower dissolved oxygen. Higher TEH loading may lead to inhibition of nitrification o r cornpetition for resources from the hydrocarbon mineralizing microbes. Higher ammonia loading may lead to inhibition of TEH mineralization. Lower dissolved oxygen could iimit nitrification. As ammonia is oxidized via nitrification approximately 4.6 g of oxygen is consumed per gram of ammonium nitrogen oxidized, which can deplete oxygen in the aquatic

environment (Reddy and Patrick, 1984). Lower d issolved oxygen could also limit hydrocar bon mineralization. The cause of decreased removal rates at higher loading was not determined.

5.2-5 1s treatment sustainable?

Although the potential impacts of oil s ands wastewater (Le., Dyke and Pond water ) on emergent aquatic macrop hytes (hereafter referred to as macrophytes) are still not well understood, there is potential that both macrophyte production and decomposition may be negatively impacted b y oil sands wastewater. A s significant differences were only observed in 1993 (under higher loads), i t is possible impacts can be controlled by controiling loading. Additionally, natural impacts such as t h e effect of m u s k r a l or beavers may significantiy affect the success of treatment constructed wetlands. A plan to address them should b e included in the design, irnplementation and operation of treatment wetlands.

T h e most significant observation is tha t macrophytes continued to grow and appeared to "thrivc" in the oil sands wastewaters after three "summers" of application. Given the high natural variability in production in wetlands, the effects of climate, muskrat activity and wastewater loading cannot b e separated. Yacrophyte production tended to b e lower in trenches receiving oil sands wastewater, though not significantly. However. the sampling design may have b e n inadequate due to relatively high variation and a low number of replicates as suggested by the a posteriori power analysis. The lack of significan t differences suggest no impact: however, the power of the s t u d y was low so the potential for accepting the nul1 hypothesis when i t w a s false was high. Larger-scale, longer-term demonstration wetlands are required before irnplementing treatrnent wetlands a s p a r t of mitigation design/mine closure.

Leaf litter provides surface area and nutr ients for bacteria, and therefore can indirectly affect contaminant removal. In addition. decomposition plays an important role in overall wetland function by recycling nutrients. Although trends in decomposition were not statistically significant in al1 s tudy years, there was some indication that oil sands wastewater could reduce litter decomposition. .4 posteriori power andys is indicated tha t tbere were insufficient replication a t longer exposure periods, while at shorter exposure periods i t was acceptable ( i . . power of 80 %).

Differences in decomposition under shor ter exposure periods were only

detected under t h e higher loading in 1993. However. t h e cause of t h e observed decreased decomposition rates could have been the h igher contaminant loads o r low dissolved oxygen o r some combination. The field s tudies did not determine cause.

The impact of reduced macrophyte production and/or decorn position rates on treatment efficiency and sustainability is unknown. In 1994, when the appa ren t production w a s s u bstantially reduced by muskrats, wastewater treatment efficiency w a s not affected (Chapter 2). While no clear t rend of contaminant impact on macrophyte production and/or decomposition was observed. i t is important to note that ioss of wetland function couid limit the use of wetlands as a reclamation option and perhaps also as a treatment option for t h e oil sands industry.

5-3 Applicability b the 0i.i Sands Industry

The oil sands i ndus t ry requires design cr i ter ia to implemen t treatment wetlands. To maintain removal rates of 70 % o r more, t he loading rates, should not exceed 270 mg m.' d-' and 210 mg rn-' d-' for amrnonia and TEH, respectively. Assuming the same water chemistry conditions as in 1994, the hydraulic loading rate should not exceed 1.8 cm/d or 0.6 cm/d to maintain ammonia and TEH removal. respectively . The estimated hydraulic ioading r a t e for TEH rernoval is Iower than for arnmonia. and therefore could limit t he design.

Aiternatively, loadings eould be reduced by some form of pre-treatment and/or removal rates could be enhanced to reduce land requirernents. Althoug h the re are some indications that phosphate addit ions and increased oxygen (by alternating pond/wetland cells) enhance treatment (Bishay and Six, 1996). fu r the r s tudy would be required to develop design criteria. Should land avaiiability not be a n issue and treatrnent wetlands a r e the desired technology then sustainability must be ensured .

Under Lighter loadings no dif'ferences in macrophyte production o r decomposition were observed. However, at higher loadings rnacrophyte production and decomposition (of selected species) may be impacted. This sugges t s t ha t a iarger-scale, longer-term larger , demonstration project is needed t o assess t h e sustainability of wetiand treatment efficiency a n d of wetland f unction prior to the im plernentation of f u i l - s a l e facilities.

5.4 Likrature Cited

Bishay, F.S. and P.G. Six. 1996. Constructed wetlands for t h e treatment of oil sands wastewater. Technical Report #5. Prepared for Suncor Inc., Oil Sands Group by EVS Consultants Ltd,, North C'ancouver. BC. 12 Chapters (236 pp.) + appendices .

Dalsgaard, T. and F. Bak. 1992. Effect of acetytene on ni t rous oxide reduction and sulfide oxidation in batch and gradient cu l tures of Thiobacillus denitrificans. Applied and En vironmen fal .VicrobiologS7 58(5):1601-1608.

FTFC (Fine Tailings Fundamentals Consortium ). 199Sa. Volume 1: Clark hot water extraction fine tailings. In: Advances in Oil Sands Tailings Research. Alberta Department of Energy, Oil Sands and Research Division, Pu blisher, Edmonton. Alberta. pp. 1-84.

FTFC (Fine Tailings Fundamentais Consortium). 1993b. Volume II: Fine tails and process water reclarnation. In: .-ldvances in Oil Sands TaXngs Research. Alberta Department of Energy. Oil S a n d s and Research Division, Publisher, Edmonton, Alberta. pp. 1-52.

Freney, J.R.. R. Leuning, J.R. Simpson, O.T. Denmead and L A . Yuirhead. 1985. Estimating ammonia volatilization from flooded r ice fields by s i m p Iif ied tech ni q ues . Soi1 Science Society of -4 merica Journal 49:1049- 1054.

Gersberg, R A . , B.V. EJkins. SIR . Lyon and C.R. Goldrnan. 1986. Role of aquatic plants in wastewater treatment by artificial wetlands. kater Research 20:363-367.

Harnmer, DA. and R.L. Knight, 1994. Designing constructed wetlands for nitrogen removal, kater Science and Technoi~g.\~ 29(4):15-27.

KIoptek 1973 and 1978 cited in Richardson and Nichois. 1983

Nix. P.C. and R.W. Yartin. 1992. Detoxification and reclamation of Suncor 's oil sand tailings ponds. Environmental Toxicology and Chemis try 7(2):171-188.

Reddy. K.R. and N.H. Pa t r i ck 1984. Sitrogen transformation and loss in f looded soils and sediments. CRC Criticai Reviews in Environmentai Control 13(4):273-309.

Reed, S.C.. R.11'. Crites and E.J. Yiddlebrooks. 1993. .Laturai systems for waste management and treatment, 'IcCraw-Hill Inc., New York.

Richardson, C.J. and D.S. N i c h o l s . 1985. Ecological analysis of wastewater management criteria in wetlands ecosystems. m: P.J. Godfrey, E.R. Kaynor, S . Pelczarsk i and J- Benforado ( eds . ), Ecologica I considera tions in weüands trea tmen t of municipal wastewa ters. Van Xostrand Reinhold Company: New York. pp. 351 -39 1.

Seitzinger, S.P., L.P. Nielsen, J. Caffrey and P.B. Christensen. 1993. Denitrification measurements in aquatic s e d i m e n t s : a com parison of three methods. Biogeochemistry 23:147-167.

Van Ostroom, A-J. and J .Y. Russell. 1994. Denitrification is constructed w a s t e w a t e r wetlands receiving high concentrations of nitrate. kater Science and Technology 29(4):744.

Water Pollution Con trol Federation ( KPCF). 1990. Satura1 systems for wastewater t r e a t m e n t . MOP FD- 16. Imperia1 Printing. S t . Joseph, Y i c h i g a n

APPENDIX A

STUDY SITE DESCRIPTIONS

Appendix A.1 Contaminated Natural Wetlands and Constructed Wetlands

The s t u d y area , which includes t h e constructed wetlands and t h e contaminated wetland, is located southwest of Pond 2 between Dyke 2 and Highway 63. This area w a s d is turbed d u r i n g ear l ier mining, but was filled in with sand by 1984. A coke filter was installed along Dyke 2 in t h e fa11 of 1984 with collection Iines to take seepage water to Ex?raction. .A 15 c m Iayer of muskeg soi1 w a s spread o v e r the area in 1983, initiating of reclamation procedures for t h e area. Sur face water began pooling in t h e area, as t h e Dyke 2 west s t r u c t u r e was being built, suggest ing that the drainage system was not working effectively. The drainage lines were excavated and new lines installed along t h e toe of Dyke 2 in 1986-87. In t h e spr ing of 1988 the a rea w a s again d i s tu rbed bÿ construction although reciamation had been completed on Dyke 2 w e s t . Drainage ditches were constructed to remove some of t h e s u r p l u s water d u e t o the high water table. Ditches (sloughs) were left in place with muskeg soi1 spread d o n g their edges. The water in these di tches is a combination of local shallow ground water, su r f ace run-off and possi bly tailings pond water seepage.

Since 1988 some of t h e surrounding areas have been colonized by emergent aquatic vegetation: consequently, prornpting a feasibiiity s t u d y for t h e u s e of constructed wetland as a treatment "faciiity" for tailings pond water seepage. Water chernistry . toxicity a n d biological analyses w e r e unde r taken in October of 1990 on t h e water flowing th rough the marsh to establish t h e extent of any "treatment" of t h e seepage water occurring. This marsh i s now referred t o as t h e "contaminated" wetland.

Although the s t u d y was composed of s ing le samples taken on a s ingle day trends from "upstream" to "downstream" were observed. Determining t h e causes of these t r ends w a s difficult d u e to t h e inability to separate the effect of dilution from the "treatment" process on t h e tailings pond water seepage in t h e s t u d y area: however. t he se resu l t s appeared promising and therefore Suncor decided to investigate t h e u se of wetland for treatment of their waste water. Yonitoring of t h i s wetland continued in the 1992 fieid season with t h e data support ing t h i s treatment of dyke seepage from inflow to outflow.

As a resul t of t h e field s t u d y in October 1990, Suncor began investigating t he potential fo r wetlands to treat dyke seepage and Pond 1.4 top water. To investigate th i s treatrnent process a constructed wetiand was built in the vicinity of t h e contaminated wetland.

T h e constructed wetlands were built in Yay and ear iy J u n e of 1991. The constructed wetland s tudy area covers approximately 1.4 ha. There are nine t renches which are 50 m long a n d 2 m wide a t t h e bottom. The t renches were planted with Typha sp. (cattai l) and Scripus sp. (bu l rush) in the week of June 10 t o 14 and allowed to establish for the remainder of t h e 1991 growing season unde r ditch water conditions. There will be three t reatments each having t h r e e r ep l imte trenches: t reatment 1 = control (ditch water), t reatment 2 = dyke seepage water, treatment 3 = Pond LA

tailings pond top water. k a t e r from t h e t h r e e treatments began flowing June 29. 1992 at rate of ' 2 L/rnin.

These t renches were fed from one of t h r e e tanks being supplied from t h e source (e.g., ditch - control , b m s t e r station - dyke seepage or Pond 1.4 top water) by ei ther vacuum t ruck (for dyke seepage and Pond 1.4 top water) o r pump through f i re hose (for control water). Residence times in t h e tanks w a s approximately one week. This experimental facility w il1 remain operational for at least 3 - 5 years.

Throughout t h e operation of these cons t ruc ted wetland t renches various sampling programs will be implemented. The design for 1992 included chernical and toxicological analyses of t h e water from inflow, mid- trench and outflow. hydrological monitoring, a n d characterization of vegetation. phytoplankton, zooplankton, and benthic and microbial populations.

Appendix A 2 Referenœ Wetlands

A11 reference wetlands w e r e not monitored in each s tudy year d u e to logistical access problems, etc. A brief description OP each wetland is provided below . Shipyard Lake

Shipyard Lake i s approximateiy 85 ha a n d is situated in the Athabasca River floodplain. This wetlands was rnonitored in 1992 only as access w a s difficult and because the wetlands quite different from the experimental wetlands. The southern portion of t h e wetlands can be classif ied as rnarsh while the northern portion i s a mixture of swamp and marsh. The vegetation a t the southern (upstream) e n d is dominated by TJ-pha latifolia (cattail). Phragmites australis ( r eed ) and -4corus calanus (swetflag). The middle zone i s characterized by Equisetum fluviatile ( water horsetail), P. australis, and T. latifblia. The nor thern (downstream) end was characterized by swamp type vegetation including: tall s h r u b s (2-10 m) Ahus tenuifolia ( r iver a lder) and Sa&- s p p. ( w illow ) . 'Yenyan thes trifoliata (buck been), T. latifdia, P. australis , Scirpus vdidus (common g rea t bulrush) , Scirpus microcarpus (small fruited bulrush). -4. calanus, Sparaganium angustîfolium ssp. emersurn f bur-reed ), h p h a r variegaturn (yellow pond-lily ) and Lernna Open water species include Elodea canadensis (Canada water weed ) and Myriophylum sp. ( water miü'oil), S. variegat um, and Lemna

Sediments had a silty-clay texture a n d had peat accumulations Iess than 50 c m in most areas.

Gra vel Pit

This reference wetiand was located approximately 15 rn from a srnall road on land reclaimed from grave1 pit operations. Aithough this wetland had a sandier subs t r a t e than the constructed wetlands i t was selected as i t was a reasonably young wetland and though t to be a better comparison than t h e well established Shipyard wetland. It w a s ais0 a smaller wetland

(approximately 0.2 ha) again potentially a be t te r comparison than the much larger Shipyard wetland. In addition. s i te access was much easier t r ip by road (Le., <30 minute t r i p by road) than shipyard wetland was by boat and portage. This wetlands was used in 1993 and 1994.

.-1OSTR-I Road

This reference wetland w a s located approximately 100 rn from t h e grave1 road to the AOSTRA facility. Althougb this wetland had a clayier s u bstrate than the constructed wetlands it w a s seiected as it w a s a reasonably young wetland and thought to be a better comparison than t h e well established Shipyard wetland. I t w a s also a smaller wetland (approxirnately 0.2 h a ) again potentially a better comparison than the much larger Shipyard wetland. In addition, s i t e access was much easier t r ip by road (i.e., <30 minute trip by road) than shipyard wetland was by boat and portage. This wetlands w a s used in 1993 and 1994.

Highway 63

This reference wetiands covered approximately 2 ha and was located about 100 m from Highway #63 and about 2 km south of the Suncor Lease. I t was monitored in 1994 only. This wetland was sought out in hopes of providing a reference wetlands with sediment more similar to the constructed wetlands compared with the Grave1 Pit or A0STR.A road reference wetlands.

APPENDIX B

WATER CHEMI STRY MmHODS AND QA/QC

Appendix B.1 - Laborabry Methods

Ammonia w a s deterrnined from a n unfiltered sample i ~ s i n g the phenolhypochiorite method of Solorzano (1969) with the prepared samples being analyzed on the Technicon AutoAnalyzer. Nitrite-nitrate was determined from a filtered sample (0.45 pm HAWP Millipore filter) using t h e cadmium-copper reduction method of Stainton et al, (1977). TKS w a s determined using the sulphuric acid-copper sulphate method of D'EIia et al. (1977) and digested samples w e r e analyzed using the Technicon AutoAnalyzer.

Suncor La bora tory

Ammonia was determined using an Orion Yodel EA 940 Expandable Ionanaiyzer fitted with an ammonia selective electrode (Corning #476130) in accordance with the method 4500 - NH3 F outlined in APHA (1989). Nitrite- nitrate nitrogen was determined using Ion Chromatography ( IC) using a carbonate biocarbonate element (1.7 mhl) through a 4 mm packed polystyrene column. TKN was not analyzed at Suncor.

.4SL La bora f ory

Ammonia was determined colorirnetrically using phenol and hypochlorite reagents in the presence of manganous salt in accordance with the method 4500 - N H 3 - A, D and F as outlined in APHA (1989). Nitrite-nitrate w a s determined colorimetrically, af ter reduction with hydrazine sulphate, using sulphanilamide and 5- 1-naph thyi-ethylenediamine dihydrochloride reagents in accordance with the method 4300 - 503 - A and H as outfined in .WHA ( 1989). T K N w a s determined using a selective electrode af ter sulphuric acid, potassium sulphate and mercuric sulphate digestion in accordance w i t h t h e method 4500 - Y(org) - Â and B as outlined in APHA (1989).

Appendix 8-2 - QA/QC

As noted in Appendix B. L th ree laboratories conducted t h e analyses. Yost of the analyses were conducted by the Cniversity of Alberta; however, initially t h e other two labs were used. Cornparisons were made between labs, although not for all sarnple periods, and t he results are provided in Tables B.2-1, B.2-2 and B.2-3. The relative percent difference for ammonia within t h e th ree labs was LI%, 3% and 11% for the Cniversity of Alberta , Suncor and ASL. The RPD was 21% between ASL and Suncor and Lniversity of Alberta and Suncor.

Cornpanson Cornparisons within L.4 2 1 within Suncor within ASL between C.4 and Suncor between ASL and Suncor within C.4 7 - within Suncor 5 within ASL .

\ - beween CrX and Suncor - i between ASL and Suncor 18

C. 1 Introduction

An unders tanding of t h e water balance o r budget is paramount to determining contaminant fate/removal in treatrnent wetlands. That is. a n accurate water mass balance is a prerequis i te to a n accurate contaminant mass balance (Kadlec and Knight, 1996). The mass balance m u s t include: 1) a weI l defined system (Le., a defined volume in space s u c h as t h e sur face water in a surface f low wetlands), 2) a time period over which inputs and ou tpu t s are totalled. 3) al1 inputs and outputs , 4 ) a n y reactions tha t produce o r consume the chemicai const i tuent of interest. The water mass balance was determined on a monthly basis for each s t u d y year, b y measuring water flow in and out, precipitation and evapotranspiration. Groundwater inputs and outputs were assumed to be negligible s ince a plastic liner w a s installed in each t rench.

The climate station installed in 1992 between the constructed wetlands and t h e natural wetlands was monitored on a daily basis. T h e station inciuded a thermistor probe and a rna.imum/minimum thermometer in a s tandard Stephenson screen (air temperature), recording and non-recording rain gauges (rainfafl dep th and intensity), a 3-cup anemometer (bind r u n ) , and a Class A evaporation pan (evaporation). Signais from t h e therrnistor probe and tipping bucket rain gauge were recorded automatically on da t a loggers (Lakewood R-X-6) at a scan interval of 10 minutes integrated eve ry 60 minutes (except in 1994 when data was integrated every 6 hours) . Yanual collections were made from the evaporation pan and anernometer as well a s a max/min thermorneter. The time and frequency of manual collections varied, b u t generally occurred in t h e morning (generallx between 0700 and 1000 h ) at least 4 days a week. Recorded data were d o m loaded to a cornputer file a t reguIar intervals. The ciimate station was fenced to prevent interference €rom animais.

Inflow and outflow rates were measured rnanually once o r t w i c e weekly a t Stations A a n d D using a bucket and a stopwatch. Inflow rates were adjusted when necessary. hater levels were also measurcd at Station D continuously using Lakewood water Ievel recorders which were placed in stilling weus adjacent to the weirs. Kater level da ta were recorded with R-X-6 data loggers.

C.3 Results and Discussion

The monthly ave rage temperature and precipitation data for t h e 1992- 1994 monitoring per iods ( Atmospheric Environment Service. 1992. 1993, 1994) compared with normal. temperature and precipitation ( Le., 30 year average data; Atrnospheric Environment Service, 1993) for t h e Fort YcYurray .4

Climate Station are given in Table C.1. Yean daily temperatures from Yaj' to September were generally cooler in 1992 and 1993 than t h e 30 year average for Fort McMurray, while the opposite w a s observed for 1994. Total monthly precipitation in the same period Kas variable (Le.. no consistent t r e n d s between years for al1 months): however. precipitation in 1992 and 1993 w a s about the sarne a s t h e 30 year average for Fort McHurray, while t he re was less precipitation in 1994. In general. 1994 was warmer and d r i e r than average.

The rnonthiy average temperature and precipitation data for 1992-1994 compared with s t u d y averages (i.e., 3 yea r s from 1992 to 1994 inclusive) for on site climate station measurements are given in Table C.2. Average mean daily temperatures were cooler at Suncor than Fort YcWurray except in September of 1992. The w a r m e s t mean daily temperatures (monthly average) were observed in July 1994 (manual 17.1 OC; datalogger 17.7"C) and t h e coolest in September 1992 (datalogger 6.Z°C)/September 1993 (manual 69°C). Extreme temperatures are provided in Table C.3a and 3b. The extreme minimum was observed in September 1992 (manual -9°C: datalogger -7.3"C) and t h e extreme maximum w a s observed in July 1994 (manual 34°C: datalogger - 34.8"C). Total monthly precipitation was about the same a t Suncor s i te compared with Fort YcYurray. On average 1992 rnay have been sl ightly wetter and 1993 and 1994 sfightly dr ier . The highest monthly precipitation was observed in J u n e 1992 (manual 119 mm: datalogger 115.2 m m ) and the lowest in August 1994 (manual 15 mm: datalogger 20.4 m m ) .

In general, the manual and datalogger collections provided sirnilar data. However, in water balance calculations t h e datalogger data was used only when a s t age discharge relationship was derived. There is uncertainty o r e r r o r associated with t h e measurernent of precipitation: I ) misreading t h e scale (Winter, 1981): 2) faifing to consider water (0.024 c m ) required t o moisten a d r y gauge (Winter, 1981); 3) failing to consider the speed a t which t h e tipping bucket tips dur ing heavy rain which can resul t in underestimates of 5% (Linsely et al.. 1958); 4 ) location of the gauge above t h e ground can resu l t in underestimates of 5 4 5 % at I rn cornpared with ground level or higher if comparing for single storm event (Neff, 1977).

Another source of error occurs during aerial averaging of point da t a (i.e., each rain gauge) over thg s tudy area. For example. ,with a gauge densi ty of 1 per 2,100 ha (21 km') or 1 per 240 ha (2.4 km') would result in a n e r r o r of 11% o r 4% respectively. However. th i s e r r o r is likely to be insignificant s i nce t h e s tudy a rea is relatively small ('3 to 3.5 h a which encornpasses the climate station, contaminated natural wetlands, t h e constructed wetlands and surrounding area).

The tipping bucket data (i.e., t h e datalogger da ta ) was chosen ove r t h e manually collected rain gauge data primarily because t h e gauge was not read eve ry day dur ing the s tudy and as such would likely have t h e higher error. In general, one can assume approximately 1 to 5% instrument e r r o r , 5 t o 15% placement e r r o r , up to 20% for Iack of wind shield e r ro r , and < 4% averaging error (for a relatively small s tudy area) (biinter. 1981). In th i s s t u d y o u r t ipping bucket rain gauge was placed at ground level. therefore e h i n a t i n g placement e r ror . However, w e did not have a n y wind

shie lds in place. The total possible error associated with t h e collection a n d use of t h e precipitation data could h a v e been as high as 25%.

Evapot ranspiration

Se t (as rneasured) and total (corrected fo r precipitation) pan evaporation (from a Class 'A* Evaporation Pan a t Suncor s i t e ) data are given in Table C.4. The monthly average evapotranspira t ion fo r 1992-94 compared with s t u d y averages (Le.. 3 y e a r s from 1992 to 1994 inclusive) fo r t h e on s i t e weather stat ion rneasurernents a n d Corn plementary Relationship Areal Evapotranspiration (CRAE) at t h e For t YcMurray climate stat ion (i.e., 17 year ave rage data: Bothe and Abraham. 1987 and 1993) are given in Table C.S. Although evaporation pan coefficients have been s h o w to Vary (Winter, 1981). a pan to wetlands coefficient of 0.8 w a s used to est imate wetlands evapotranspiration ( Kadlec, 1989).

Evapotranspiration, a s determined from pan evaporation, generally increased from 1992 to 1994 in a n y given month. Again indicating t h a t 1994 was t h e dr ies t year in t h e t h r e e year s t u d y - Although on ave rage J u a e tended to have the highest evapotranspira t ion, it was not t h e highest in =ll years. Evapotranspiration was lowest in September 1992 (23 mm) a n d highest in J u n e 1994 (108 mm).

Hydraulic Reten Lion Tjme

I t i s pararnount t o characterize t h e interna1 wetland water flows because s t ud i e s of contaminant removal/nutrient cyciing could otherwise b e anomalous ( H a m m e r and Kadlec, 1986). The design flow (i.e., t h e flow designated at t h e beginning of each s t u d y season) and therefore hydraul ic retention time (HRT) varied between s t u d ÿ years . The design flows were 2 L/rnin in 1992 and 1994 and 6 L/min in 1993. The nominal detention time can be calculated by dividing t he volume of t h e tre ch bÿ t he input flow 4 rate. The estimated volume of each t r ench i s 52.5 m (i.e., 50 rn long x 3.3 m mean width x 0.35 m mean depth x a n est imate of porosity, 0.9). a n d the re fore t he nominal detention time (or hydraul ic retention time) fo r 2 L/min and 6 L/min flow was 19.1 d a y s a n d 6.4 days, respectively; assuming approxirnate plug-flow conditions. Approximate plug-flow conditions d o not generally occur in free-water s u r f a c e wetlands, even in long narrow wetlands (Kadlec et al., 1993). Experimental verif ication of these estimated HRTs w a s required to accurately p red ic t t h e range of contact tirnes.

In 1993. hydrologie s tudies were conducted us ing bromide a s a tracer to determine t h e HRT at 2 , 6 and 20 L/min. The median HRTs were 17.1, 5.3 and 1.3 days, respectively (Nix et al,. 1994). The median HRT r ep re sen t s t h e time it took 50% of t h e recovered bromide to pass through the const ructed wetlands trench, whife t h e 10% HRT a n d 90% HRT re fe r to t h e t i m e i t takes 10% and 90% of t h e bromide to pa s s through t h e trench. The 10% HRTs fo r 2 and 6 L/min flow were 10.8 a n d 3.6 days, respectively; whiie t h e 90% HRTs were 21 and 8.3 days, respectively (Nix et al., 1994). The median HRTs were actually v e r y similar to t h e nominal HRTs; and there fore a 6 day HRT was assumed for t h e 6 L/min flow rates in al1 t renches.

In 1994, hydrologie s tudies included both bromide t racer and rhodamine dye tracers. These s tudies were only conducted in t h e spli t trenches a s i t was assumed that HRTs would not change from year t o year assuming the flows tested were t h e same as those tested in 1993. The rhodamine dye was used primarily to assess tenplast berm integrity: while the bromide wzs used for HRT determinations and berm integrity. Based on flow, bromide t racer and rhodamine dye data, berm integritÿ was questionable for t rench 3 (T3) and trench 5 (TS), therefore inflows and outf1ows were combined for t h e purposes of contaminant removal. Results for trench 7 (T7) berm were not a s obvious: that is, some leakage was indicated by t h e rhodamine dye, but not t he flow o r bromide. However, t h e east and west sides were also combined for T7 for the pürpûses of the contaminant removal. The HRTs w e r e 3 days for T3 (12 L/min), 24 days for T5 (1.5 L/min) and 17 days for T7 (2.25) o r 9 days for T7W (2 L/min in half of t rench) and 72 days for T7E (0.5 L/min in haif of t rench) (Nix et ai., 1995). The HRTs for t h e remaining trenches w a s assumed to be 18 days at fiow ra t e of 2 L/min based on the mean of 1993 tracer data and t h e nominal HRT.

k'a ter Balance

Yean daily inflow measurements were made by s top watch and bucket. Yean daily discharge (Le.. outflow, L/min) was calculated by linear regression of instantaneous discharge measurements ( b y s top watch and bucket) on stage measurements ( b y float Ievel indicator and datalogger) in addition to just means of manual discharge measurements.

Cu'ater balance estimates are valuable since they can be used to assess the role of precipitation and evapotranspiration on flow through t h e wetlands. Precipitation can increase flows and storage, while evapotranspiration may reduce flows and concentrate contaminants. Transpiration i s the process where water is drawn from plant roots and transported to stomata in the leaves where the water is finally evaporated. Thus transpiration is the "engine" behind plant uptake processes. Evapotranspiration w a s estimated by t h e evaporation in a class 'A' pan monitored a t the s t u d y site, and together with precipitation and flow data, water budgets for each trench w e r e determined.

Flows for sampling days were extrapolated from flows around the sampling days. Average monthly flows are also given for each trench. Lsing monthly flows, precipitation and evaporation data the overall water budget for each trench w a s determined (Table C.6).

On average, unaccounted flow in 1992 was indicative of a n unidentified input (i.e., 7% of t h e mean inflow of 1.52 L/min came from unmeasured inputs) while in 1993 and 1994 i t w a s indicative of a n unidentified output (16% of mean inflow of 5.65 L/min and 11% of mean inf'low of 2.3 L/rnin, respectively left from unmeasured outputs). In al1 years t h e unaccounted portion of the inflow w a s less than 25% for a11 trenches except T l in 1992 and 1994 and t h e spli t t renches (Le., T3, T5 and T7) in 1994. The unaccounted flows tha t w e r e less than 25% of the inflow were likely within t h e e r r o r of t he measurement techniques and therefore i t is unlikely that

t h e t renches leaked dur ing the s t u d y period. Only T l may have leaked, and s ince i t w a s a Control t rench i t does not affect the nitrogen budget deterrnined fo r wetlands receiving high ammonia loads. Therefore, no adjustments were made to trench flows and i t was assumed that trenches were sealed from ground water inputs/outputs.

A considerable proportion of t h e inflow water is not accounted for by the outflow, precipitation. evapotranspiration, w hich sugges t s tha t t h e trenches leak. However, the methods to der ive estimates of fiow , evapotranspiration, etc. are somewhat crude. For exampte, flow measurements were made with a bucket and stop watch and only dur ing t h e day, the reby underestirnating the actual outflow, because t h e outflow is expected t o be kigher at night when evapotranspiration is lower. Stage level w a s recorded using float level and datalogger; however, at th i s time s t age flow relationships are not available fo r al1 years and therefore th i s data was not used.

The error associated with t h e various hydrometeorlogical parameters is as follows (Winter, 1981):

precipitation measured with recording rain gauge - 2% evapotranspiration measured with Class A pan - 10% evaporation pan to lake/wetlands coefficient - 15% inflow with cu r r en t meter - 5% outflow with c u r r e n t meter - 5% outflow stage discharge relationship - 5-10%

These errors may underestirnate t h e actual error in t h e case of f low rneasurements since bucket and stopwatch measurements were made. They may also underestimate error associated with evapotranspiration since the pan to lake coefficient may have more error due to the small size of t h e wetlands which may result in increased evapotranspiration d u e to t h e oasis effect (Morton, 1983) and wicking. These errors would be the equivalent to t h e following on average from J u n e to September inflow ( u n i t s in L/rnin):

total precipitation e r r o r of 2% --> 0.004 L/min total evapotranspiration error of 25% --> 0.07 L/min total inflow e r r o r of 5% --> 0.08 ( for 1.52). 0.28 (for 3-65) , 0.1 (for 2.33) L/min total outflcw e r r o r of 15% --> 0.35 (for 2.33), 0.68 (for 4 . 3 , 0.32 (for 2-15) L/min

Therefore, t he total error mu id have b e n 0.5, 1, 0.5 L/min for 1992, 1993 and 1994, respectively. These errors are the equivalent to 33%, 18%, 21% of the inflow for 1992, 1993 and 1994, respectively. In 1992, t h e only t rench where t he unaccounted flow exceeded t h e measurement error of 33% was T l (-72%). while ail the r e s t were s 15% of t h e inflow. In 1993, t he unaccounted flow exceeded t h e measurement error of 18% in T l (24%). T4 @O%), T5 (24%) and T8 (21%). In 1994, t h e unaccounted flow exceed t h e measurement error of 21% in Tl (35%) and the t h r e e spli t t renches (-40%, - 40% and -29% for T3, T5, T7, respectively). Essentially, T l outputs were overestimated or inputs underestirnated in 1992 which is not indicative of a leak: however, in 1993 and 1994 t h e r eve r se was t r u e therefore indicating

a leakage may have occurred. Although the re w e r e other t renches where t h e unaccounted flow exceeded t h e estimated e r ror , these values were very close and therefore no other trench's integrity w a s questioned except for t h e split trenches in 1994. The integrity of Tl does not affect t h e determination of t he nitrogen budget as i t was Control trench.

Table C.1 Average temperatures and precipitation rates at the Fort McMurray A Climate Station for 1992-1994 compared with 30 year averages.

May

J u n e

J u l y

S e p t e m ber

na not available

Mean Daily Temperature ("cl

Total kionthly Precipitation (mm )

Sonna1 ( 1961 - 1990)

10.1

14.6

16.6

15.2

9.1

1992

9.2

14.1

15.7

14

6.7

Kormal ( 1961-

1990 1 l

40.7

63.9

79.1 ,

71.8

51.4

1992

. 20.8

108.2

39.6

52.4

69.0

1993

na

13.4

15.3

14.5

8.6

1993

na

72.8

78.6

5 0 .

34.4

1994

11.2

15.2

17.8

16.3

12.0

1994

22.8

61.0

94.6

7.4

19.0

Table C.2a Average temperatures and precipitation ra tes at the Suncor Climate Station for 1992- 19% compared with 1992-1994 averages: Manual,

- incomplete data s e t

Yonth

Yay

June

Ju ly

August

September

Table C.2b Average temperatures and precipitation rates at the Suncor Cl imate Station for 1992-1994 compared with 1992- 1994 averages: Datalog ger .

Month Mean Daily Temperature

Yean Daily Temperature ( O c )

June 12.5 13.4 14.9 13.6

Total Lionthly Precipitation ( m m )

August 1 12.9 1 14.2 1 16.0 1 14.4 I 1 I I

1992

-

119

40

42

55

1992

- 12.9

14.5

13.2

7.4

Total Monthly Precipibtion (mm)

1991

10.4

14.4

17.1

15.9

11.0

1993

11.3

12.3

13.4

13.1

6.9

( 1 9 9 2 - 9 4 )

10.9

13.2

15.0

14.1

8.4

( L9!3Z-!i4I

33

91

59

29

4 1

1993 1994

47

67

81

29

33

20

86

57

15

36

Table C.3a Minimum and maximum daily temperatures at the Suncor Cl imate Station for 1992-1994 cornpared with 1992- 1994 averages:

Yinimum Daily Temperature ( O C )

l

Yaximum daily l Temperature ( OC)

Yay

June

July

August

Septern ber

Table C.3b Minimum and maximum daily temperatures a t the Suncor Climate Station 1992- 1994 compared with 1992- 1994 averages: Datalog ger .

Maximum daily Temperature ("0 I Minimum Daily

Temperature (OC)

June 1 - 1 -2.5 1 -1.9 1 -2.0

August

Septem ber

Table C.4 Net and Total Evaporation (using a Class 'A' Evaporation Pan) at the Suncor Wetiands Study site.

Month

May

June

July

August

Septem ber

Monthly Net ~vaporationl ( m m )

Monthly ota1 T Evaporation (mm)

a b

- net evaporation was the change measured in the evaporation pan - total evaporation equals net evaporation + precipitation

Table C.5 Evapotranspiration at the Suncor Wetlands S tudy site for 1992-1994 compared with 17 year averages at Fort Ilc-iurray. AB.

a 80% of total evaporation is generally accepted to equal evapotranspiration in a wetlands (Kadlec and Knight, 1996)

Lionth

P

Yay

Ju ne

July

August

Sep tem ber

~ v a ~ o t r a n s ~ i r a t i o n ~ (mm)

Normal ( 1972-1992)

I

39

66

80

54

16

Three ( 1992-94) year

Yonthly

1994 1992 1993

-

53

73

61

23

67

97

75

67

30

81

108

86

81

4 0

74

86

78

70

31

NOTE TO USERS

Page(s) not included in the original manuscript are unavailable from the author or university. The manuscript

was microfilmed as received.

UMI

C.4 References Cited

Atmospheric Environment Service. 1992, 1993, f 994. Slonthly climatologicaJ summary, For t McMurray A ( Y M M ) Alberta. For t McMurray Weather Office. Atrnospheric Environment Service, Environment Canada.

Atmospheric Environment Service. 1993. Canadian ciimate normals, 1961 - 1990; Prair ie Provinces. Canadian Climate Program. Environment Canada. Ottawa. 266 pp.

Bothe, R.A. and C. Abraham. 1993. Evaporation a n d evapotranspirat ion in Alberta 1986-1992: Addendum. Hydrology Branch, Technical Services Division, Water Resource Management Services, Alberta Environment. 3 pp. + apps .

Bothe, R.A. and C. Abraham. 1987. Evaporation a n d evapotranspirat ion in Ai berta 1912- 198% Hydrology Branch, Tech nical Services Division, Water Resource Management Services, Ai be r t a Environment. 21 pp. + apps.

Hammer, D.E. and R.H. Kadlec. 1986. A mode1 fo r wetland sur face water dynarnics, Water Resources Research. 22(13):1951-1958.

Kadlec, R.H. 1989. Chapter 3: Hydrologic fac tors in wetland water treatrnent. In: D.A. Hammer (ed. ). Constructed wetlands for wasfewater treatment: municipal, industrial and agriculfural. pp. 21- 40.

Kadlec, R.H. and R.L. Knight. 1996. Treatment wetlands. Boca Raton: Lewis Publishers. 893 pp.

Kadlec, R,H., h:. Bastiaens, a n d D.T. Grban. 1993. Hydrological design of f ree water su r face t rea tment wetlands. In: C.A. Yoshiri (Ed.) Constructed weflands for water quaiity improvement. Boca Raton, FL: L e w i s Publishers. pp, 77-86.

Linsely, R.K., K A . Kohler and J.L.H. Paulhus. 1958. H-vdrology for engineers. New York, hY: McGraw-Hill. 340 pp.

Yorton, F.I. 1983. Operational estimates of a rea l evapotranspiration a n d their significance to t h e science and practice of hydrology. Journal of Hydr01og.v. 66:1-76.

'leff, E.L. 1977. How much ra in does a rain gage gage? Journal of Hydrology. 35(3/4):213-220.

Nix. P.G., S.H. Hamilton, L. Bendeil-Young, C.P. Gunter , F.S. Bishay and M.D. Paine. 1994. Constructed wetlands for t h e treatment of oil s a n d s wastewater: Technical Report #3. Prepared fo r Suncor Inc. - 0i1 Sands Croup. EVS Consultants , North Vancouver, BC. 222 pp. + appendices.

Six, P.G., F.S. Bishay, S.H- Hamilton and Y.D. Paine. 1995. C o n s t r u c t e d wetlands for the treatrnent of oil sands w a s t e w a t e r : Technical Report #4. Prepared for Suncor Inc. - Oil Sands Croup. EVS Consultants, Korth Vancouver, BC. 386 pp. + appendices.

luinter, T.C. 1981. Cncertainties i n e s t i m a t i n g the water balance of Iakes. Water Resources Bulletin. 17(1):82-115.

APPEXDIX D

TREATYENT EFFICIENCY COMPARISOS KITH THREE YODELS

Appendix D.1 Liodel Descriptions

Yodel 1 assumes tha t constructed wetlands are attached g r o ~ t h biological reactors and tha t the i r performance can be estimated by f i rs t -order plug- flow kinetics ( R e d et al., 1995). T h e basic equation i s a s f'ollows:

Where: Co = initial/inflow ammonia concentration (mg N/L) C, = effluent/outflov a m m o n i a concentration (mg N/L)

2 A, = wetland area (rn ) kt = temperature dependent first order reaction rate constant. daYsml

For: 0°C k t = 0 (d-I) 1- 10°C (T-10'1 k t = 0.1367*1.15 , (d-'-4 >lO°C k t = 0.2187*1.048~~-'@ ' ( d )

y = rnean water depth ( m ) n = porosity or space available for water to flox through t h e wetland Q = Tlow ( m 3 / d ) HRT = hydraulic retention time ( d )

Mode1 2 is based on a regression analysis of' data from constructed a n d natural wetland treatment systems (WPCF, 1990). T h e basic equation is a s follows:

In Ce = 0.688 ln Co + 0.655 ln HLR - 1.107

Nhere: Co = initial/infIow amrnonia concentration (mg N/L) C = effluent/oufflow ammonia concentration ( m g S/L) HLR = hydraulic loadm g rate (cm/d Y 1 i.e., flow ( m / d ) * area ( m ) * 100

Yodel 3 is based on a regression analysis of data from 17 free surface constructed wetland treatment systems (Hammer and Knigh t , 199.1 j. The basic equation is as follows:

W here:

L, = ammonia loading ( k g S/ha/day) i.e.. Cd 1000*Q/A/100 Co = initial/'nflow ammonia concentration (mg S/L) 1 Q = flow (m /d)

2 -4 = wetland area ( m ) C, = effluent/outflow ammonia concentration (mg S/L)

Merences Cited

H a m m e r , DA. and R.L. Knight. 1994. Designing constructed wetlands for nitrogen removal. kater Science and Te~hnolog~v. 29(4):15-27.

Reed, S.C., R.K. Crites and E.J. Middlebrooks. 1993. .Vatural systems for waste management and treatment. McGraw-Hill Inc., Xew York.

kater Pollution Con trol Federation ( W C F ) . 1990. Satura1 systems for wastewater treatment. YOP FD- 16. Imperia1 Printing, St. Joseph. Michigan.

Appendix 1.2 Predicted Results frorn the Three Model

Date HRT a - dean

lnflow Param ters 1 Ouiflow Parameters

NH4 Flow Load ~emperatureJ NH4 Flow Load Temperature Kt mode11

0.2117 0.00 0.0193 0.00 9,0000 9.00

O 1722 O01 01694 001 O 1706 O01 02067 O00 O 2087 O 00 O 2072 O 01

01726 001 01822 OOC O1800 O01 01860 OOC O 1941 001 01950 OOC 02005 00C 02126 00C O 2151 OOC O 2048 O01 O 1996 O O[ O 1954 OC11 01860 001 01874 00 ' 01847 00 ' 01931 OCK O 1817 00 ' O1698 00 ' 02015 00( O 1945 00 ' 01843 001 O 1639 0 0 ' 01605 0 0 ' O1550 00'

Flow

> m m o m m r ~ ) m m m m m m m ~ n m m m ~ ~ w a m ~ ~ ~ r l ) < o < ~ ~ m < ~ f i o < ~ 3 (ON q m m k N h N b b N & m

Q t - ' m r c - " < O u? Q - O m.-

<L; V m m r -

C (00 w < o m

m m œ e c

c a a a a a a a a a a a a a a a a a a a a a a a a a a a a a a a a

O - r O O ~ ~ - r r O r r r r - ~ - 0 0 r r ( D C O O O O O O O Q O O O O O O O O O O o o o c y ~ e O O O O O O O O O O O O O O O O ~ ~ 6 0 0 0 r q

TEST TARGET (QA-3)

y &- /i'

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