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Produced Water: Overview of Composition, Fates, and Effects

Authors:
  • Battelle Memorial Institute, Duxbury, MA
  • Kenneth Lee Research Limited

Abstract and Figures

Produced water (formation and injected water containing production chemicals) represents the largest volume waste stream in oil and gas production operations on most offshore platforms. In 2003, an estimated 667 million metric tons (about 800 million m3) of produced water were discharged to the ocean from offshore facilities throughout the world. There is considerable concern about the ocean disposal of produced water, because of the potential danger of chronic ecological harm. Produced water is a complex mixture of dissolved and particulate organic and inorganic chemicals in water that ranges from essentially freshwater to concentrated saline brine. The most abundant organic chemicals in most produced waters are water-soluble low molecular weight organic acids and monocyclic aromatic hydrocarbons. Concentrations of total PAH and higher molecular weight alkyl phenols, the main toxicants in produced water, typically range from about 0.040 to about 3mg/L. The metals most frequently present in produced water at elevated concentrations, relative to those in seawater, include barium, iron, manganese, mercury, and zinc. Upon discharge to the ocean, produced water dilutes rapidly, often by 100-fold or more within 100 m of the discharge. The chemicals of greatest environmental concern in produced water, because their concentrations may be high enough to cause bioaccumulation and toxicity, include aromatic hydrocarbons, some alkylphenols, and a few metals. Marine animals near a produced water discharge may bioaccumulate metals, phenols, and hydrocarbons from the ambient water, their food, or bottom sediments. The general consensus of the International Produced Water Conference was that any effects of produced water on individual offshore production sites are likely to be minor. However, unresolved questions regarding aspects of produced water composition and its fate and potential effects on the ecosystem remain. Multidisciplinary scientific studies are needed under an ecosystem-based management (EBM) approach to provide information on the environmental fates (dispersion, precipitation, biological and abiotic transformation) and effects of chronic, low-level exposures to the different chemicals in produced water.
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Produced Water: Overview of Composition,
Fates and Effects
Jerry M. Neff1, Kenneth Lee2, Elisabeth M. DeBlois3
1Neff & Associates LLC, 20 Templewood Dr., Duxbury MA, 02332, USA
2Centre for Offshore Oil, Gas and Energy Research, Fisheries and Oceans Canada, P.O. Box
1006, Dartmouth NS, Canada B2Y 4A2
3Elisabeth DeBlois Inc., St. John’s NL, Canada
1 Introduction
Produced water often is generated during the production of oil and gas from onshore
and offshore wells. Formation water is seawater or fresh water that has been trapped
for millions of years with oil and natural gas in a geologic reservoir consisting of a
porous sedimentary rock formation between layers of impermeable rock within the
earth’s crust (Collins, 1975). When a hydrocarbon reservoir is penetrated by a well,
the produced fluids may contain this formation water, in addition to the oil, natural
gas, and/or gas liquids. Fresh water, brine/seawater, and production chemicals
sometimes are injected into a reservoir to enhance both recovery rates and the safety
of operations and these surface waters and chemicals sometimes penetrate to the
production zone and are recovered with oil and gas during production (Neff, 2002;
Veil et al., 2004). Produced water (formation and injected water containing
production chemicals) represents the largest volume waste stream in oil and gas
production operations on most offshore platforms (Stephenson, 1991; Krause, 1995).
Produced water may account for 80% of the wastes and residuals produced from
natural gas production operations (McCormack et al., 2001).
The ratio of produced water to oil equivalents (WOR) or the ratio of water to gas
(WGR) produced from a well varies widely from essentially zero to more than 50
(98% water and 2% oil). The WGR usually is higher than the WOR. The average
worldwide WOR is about 2 to 3 (Veil, this volume). The volume of produced water
generated usually increases as oil and gas production decreases (WOR and WGR
increase) with the age of the well (Henderson et al., 1999). In nearly depleted fields,
production may be 98% produced water and 2% fossil fuel (Stephenson, 1992; Shaw
et al., 1999). Mean WOR and WGR for oil and gas production from Federal offshore
waters (>4.8 km from shore) of the U.S. are 1.04 and 86, respectively (Clark and
Veil, 2009). On the Canadian East Coast, the average WOR was 2 during the life of
the Cohasset oil field (1992-1999), the first offshore production in Atlantic Canada
(Ayers and Parker, 2001). Oil and gas production from the Hibernia field on the
Grand Banks is relatively dry, with a WOR of about 1 in September 2007 (Reuters,
2007). The gas is reinjected for enhanced oil recovery.
In 2003, an estimated 667 million metric tons (about 800 million m3) of produced
water were discharged offshore throughout the world, including 21.1 million tons to
offshore waters of North America, mostly the U.S. Gulf of Mexico, and 358-
419 million tons to offshore waters of Europe, mostly the North Sea (OGP, 2004;
Garland 2005). These are underestimates of actual discharges, because reporting of
production to OGP (2004) ranged from 11 percent to 99 percent in the seven regions
of the world monitored.
For example, the estimated total volume of produced water generated in U.S.
Federal offshore waters in 2007 during production of 75.7 million m3 (476 million
barrels) of oil and 2.8 billion ft3 of natural gas was 93.4 million m3, or 256,000
m3/day (Clark and Veil 2009). About 22,000 m3/day of this produced water was
reinjected for enhanced recovery or disposal, and about 234,000 m3/day was treated
and discharged to the ocean.
Produced water production on the Norwegian continental shelf was 135 million
m3 in 2009, a reduction of abut 10% from 173 million m3 produced in 2008 (KLIF
2010). About 85% of the water was discharged to the ocean each year, the
remainder was injected.
Off the coast of Atlantic Canada, produced water discharge from the Hibernia field
increased from 17,000 m3/day in July 2007 to 20,300 m3/day in September 2007 as
oil production declined (Reuters, 2007). In 2009, the Venture field on the Canadian
Scotian Shelf was discharging 100-600 m3/day of produced water (personal
communication, ExxonMobil).
There is considerable concern over the ocean disposal of produced water from
production operations, because discharge is continuous during production, discharge
volumes are increasing in most mature offshore production areas, and the masses of
many potentially toxic organic compounds and metals are higher in treated produce
water than in the receiving waters, raising concerns about chronic ecological harm.
2 Chemical Composition of Produced Water
Produced water is a complex mixture of dissolved and particulate organic and
inorganic chemicals. The physical and chemical properties of produced water vary
widely depending on the geologic age, depth, and geochemistry of the hydrocarbon-
bearing formation, as well as the chemical composition of the oil and gas phases in
the reservoir, and production chemicals added to the production. Because no two
produced waters are alike, region specific studies are needed to address the
environmental risks from its discharge.
2
Produced water contains a variety of naturally occurring compounds that were
dissolved or dispersed from the geologic formations and migration pathways in
which the produced water resided for millions of years. These chemicals include
inorganic salts, metals, radioisotopes, and a wide variety of organic chemicals,
primarily hydrocarbons.
2.1 Salinity and Inorganic Ions
The salt concentration (salinity) of produced water may range from a few parts per
thousand (‰) to that of a saturated brine (~300‰), compared to a salinity of 32 to
36‰ for seawater (Rittenhouse et al., 1969; Large, 1990; Table 1). Most produced
waters have salinities greater than that of seawater and, therefore, are denser than
seawater (Collins, 1975). Hibernia produced water has a salinity of 46 to 195‰
(Ayers and Parker, 2001).
Produced water contains the same salts as seawater, with sodium and chloride the
most abundant ions (Table 1). The most abundant inorganic ions in high-salinity
produced water are, in order of relative abundance in produced water, sodium,
chloride, calcium, magnesium, potassium, sulfate, bromide, bicarbonate, and iodide.
Concentration ratios of many of these ions are different in seawater and produced
water, possibly contributing to the aquatic toxicity of produced water (Pillard et al.,
1996).
Sulfate and sulfide concentrations usually are low, allowing barium and other
elements that form insoluble sulfates and sulfides to be present in solution at high
concentrations. Produced water from sour oil/gas wells may contain high
concentrations of sulfide and elemental sulfur. For example, produced water from an
offshore California well contained 48 to 216 mg/L sulfide and 0.6 to 42 mg/L sulfur
(Witter and Jones, 1999). If seawater, that naturally contains a high concentration of
sulfate ( 2712 mg/L), is injected into the formation to enhance oil and gas recovery
and mixes with the formation water, barium and calcium may precipitate as scale in
the production pipes. Any radium radioisotopes in the produced water co-precipitate
with the barium scale. Some Brazilian offshore produced waters contain more than
2000 mg/L sulfate, a concentration high enough to promote barium and calcium
scale formation (Gabardo et al., this volume). Hibernia produced water, recovered
from a reservoir on the Grand Banks of Newfoundland, Canada, in 2000, contained
248 to 339 mg/L SO4, low enough to reduce the likelihood of producing large
amounts of barium and calcium scale (Ayers and Parker, 2001).
Table 1 Concentrations (mg/kg or parts per million) of several elements and
inorganic ions in produced waters of different geologic ages compared with average
concentrations in 35‰ seawater (Collins, 1975)
Element/Ion Seawater Produced Water
3
Highest
Concentration (Age1)
Range of Mean
Concentrations
Salinity 35,000 --- <5000 - >300,000,000
Sodium 10,760 120,000 (J) 23,000 – 57,300
Chloride 19,353 270,000 (P) 46,100 – 141,000
Calcium 416 205,000 (P) 2530 – 25,800
Magnesium 1294 26,000 (D) 530 – 4300
Potassium 387 11,600 (D) 130 – 3100
Sulfate 2712 8400 (T) 210 – 1170
Bromide 87 6000 (J) 46 – 1200
Strontium 0.008 4500 (P) 7 – 1000
Ammonium --- 3300 (P) 23 – 300
Bicarbonate 142 3600 (T) 77 – 560
Iodide 167 1410 (P) 3 – 210
Boron 4.45 450 (T) 8 – 40
Carbonate --- 450 (M) 30 – 450
Lithium 0.17 400 (J) 3 – 50
1 D, Devonian, J, Jurassic, M, Mississippian, P, Pennsylvanian, T, Tertiary.
Ammonium ion may be present in some produced waters at elevated concentrations,
possibly eliciting inhibitory (toxic) and/or stimulatory (e.g. eutrophication)
responses from resident biota (Anderson et al. 2000, Yeats et al., this volume).
Hibernia produced water contains about 11 mg/L NH3 (Yeats et al., this volume).
Brazilian produced water contains 22 to 800 mg/L ammonia (Gabardo et al., this
volume). However, concentrations of nitrate and phosphate often are low in
produced waters (Hibernia produced water contains about 0.35 mg/L P and 0.02
mg/L NO3), decreasing the likelihood of eutrophication in the receiving waters
(Johnsen et al., 2004).
A large zone of hypoxic (dissolved oxygen <2.0 mg/L) water develops in nearshore
bottom water over an area of more than 17,000 km2 off Louisiana each summer
(Rabalais, 2005; Veil et al., 2005; Bierman et al., 2007). The hypoxia is caused
primarily by the discharge of large volumes of water containing high concentrations
of primary nutrients from the Mississippi River. However, as there are
approximately 287 oil and gas production platforms in the hypoxic area (Veil et al.,
2005), many of which discharge treated produced water, a comprehensive
monitoring program was performed in this area to determine if the discharged
production water (~81,000 m3/day) was contributing significant amounts of nutrients
to the Louisiana nearshore waters (Veil et al., 2005; Bierman et al., 2007). Produced
water from 50 platforms, discharging ~280,000 m3/day to the hypoxic zone, was
analyzed for nutrients (Table 2). Produced water from most gas platforms contained
higher concentrations of BOD and all nutrients except ammonia. The ratio of
estimated annual nutrient loading from all the platforms in the hypoxic zone to the
annual nutrient loading from the Mississippi River ranged from 0.00003 for nitrate
to 0.07 for ammonia. The investigators concluded that produced water discharges
contributed very little of the organic loading contributing to oxygen depletion in
4
bottom waters of the hypoxic zone off Louisiana (Rababalis, 2005; Bierman et al.,
2007).
Table 2 Mean biological oxygen demand and concentrations (mg/L) of several
primary nutrients in produced water from 50 platforms discharging to the hypoxic
zone in the Gulf of Mexico off the coast of Louisiana; mass loadings are
concentration × discharge volume in kg/day (from Veil et al., 2005; Bierman et al.,
2007)
Parameter Mostly Oil Mostly Gas Oil and Gas Mass Loading
No. Platforms 6 20 24 50
Biological oxygen
demand (BOD)
595 1444 642 16,330
Total organic carbon
(TOC)
551 888 297 6400
Nitrate (NO3
-) 1.14 2.71 1.94 31.0
Nitrite (NO2
-) 0.05 0.05 0.05 1.40
Ammonia (NH4) 92 57 85 2160
Orthophospate (PO4
3-) 0.34 0.61 0.30 10.2
Total phosphorous 0.62 0.86 0.61 17.0
2.2 Total Organic Carbon
The concentration of total organic carbon (TOC) in produced water ranges from less
than 0.1 to more than 11,000 mg/L and is highly variable from one well to another
(Tables 2 and 3). Produced water from Hibernia has a TOC concentration of
approximately 300 mg/L (Ayers and Parker, 2001). Produced water from wells off
Louisiana contain 67 to 620 mg/L dissolved TOC (DOC) and 5 to 127 mg/L
particulate TOC (POC) (Veil et al., 2005). A large fraction of the DOC may be in
colloidal suspension (Means et al., 1989).
Table 3 Concentration ranges (mg/L or parts per million) of several classes of
naturally-occurring organic chemicals in produced water world-wide (from Neff,
2002)
Chemical Class Concentration Range
Total organic carbon 0.1 – >11,000
Total organic acids 0.001 – 10,000
Total saturated hydrocarbons 17 – 30
Total benzene, toluene, ethylbenzene, and xylenes (BTEX) 0.068 – 578
Total polycyclic aromatic hydrocarbons (PAH) 0.04 – 3.0
Total steranes/triterpanes 0.14 – 0.175
Ketones 1.0 – 2.0
Total phenols (primarily C0-C5-phenols) 0.4 – 23
5
2.3 Organic Acids
The organic acids in produced water are mono- and di-carboxylic acids (-COOH) of
saturated (aliphatic) and aromatic hydrocarbons. Much of the TOC in produced
water consists of a mixture of low molecular weight carboxylic acids, such as
formic, acetic, propanoic, butanoic, pentanoic, and hexanoic acids (Somerville et al.,
1987; Means and Hubbard, 1987; Barth, 1991; Røe Utvik, 1999; Table 3). The most
abundant organic acid usually is formic or acetic acid and abundance typically
decreases with increasing molecular weight (Fisher, 1987; MacGowan and Surdam,
1988; Table 4). Strømgren et al. (1995) found 43 to 817 mg/L total C1 through C5
organic acids and 0.04 to 0.5 mg/L total C8 through C17 organic acids in three
samples of North Sea produced water. Several samples of produced water from
North Sea, U.S. Gulf of Mexico, and California platforms contained 60 to 7100
mg/L total low molecular weight aliphatic organic acids (Table 4, MacGowan and
Surdam, 1988; Jacobs et al., 1992; Flynn et al., 1995; Røe Utvik, 1999). Small
amounts of aromatic acids also may be present in produced water (Rabalais et al.,
1991; Barman Skaare et al., 2007). Produced water from coastal waters of Louisiana
contained low concentrations of aliphatic and aromatic acids (Table 5). Aliphatic
acids were more abundant than benzoic and methylbenzoic acids (Rabalais et al.,
1991).
Table 4 Concentrations (mg/L = ppm) of low molecular weight organic acids in
produced water from four production facilities on the Norwegian continental shelf
(Røe Utvik, 1999), in the Gulf of Mexico off the Texas and Louisiana coast, and in
the Santa Maria Basin off the California coast (MacGowan and Surdam, 1988)
Organic Acid Formula Offshore USA Norwegian North Sea
Formic acid CHOOH ND - 68 26 - 584
Acetic acid CH3COOH 8 – 5735 Not determined
Propanoic acid CH3CH2COOH ND – 4400 36 - 98
Butanoic acid CH3(CH2)2COOH ND - 44 ND - 46
Pentanoic acid CH3(CH2)3COOH ND - 24 ND - 33
Hexanoic acid CH3(CH2)4COOH Not determined ND
Oxalic acid COOHCOOH ND – 495 Not determined
Malonic acid CH2(COOH)2 ND – 1540 Not determined
Total measured
organic acids --- 98 - 7160 62 - 761
Table 5 Range of concentrations (mg/L) of aliphatic and aromatic acids in
produced water from seven treatment facilities in coastal Louisiana (from Rabalais
et al., 1991)
Chemical Pass Furchon Bayou Rigoud 5 Other Facilities
Aliphatic acids 8.5 – 120 1.8 – 78.0 7.9 – 75.0
Benzoic acid 0.92 – 15.0 0.13 – 16.0 1.2 – 13.0
C1-Benzoic acid 1.6 – 11.0 0.089 – 14.0 1.6 – 16.0
C2-Benzoic acid 0.42 – 2.3 0.043 – 2.7 0.29 – 3.8
6
These low molecular weight organic acids are readily biosynthesized and
biodegraded by bacteria, fungi, and plants, and so represent nutrients for phyto- and
zoo-plankton growth. Organic acids are produced by hydrous pyrolysis or microbial
degradation of hydrocarbons in the hydrocarbon-bearing formation (Borgund and
Barth, 1994; Tomczyk et al., 2001; Barman Skaare et al., 2007).
Many crude oils, particularly those that have been biodegraded in the formation,
contain high concentrations of naphthenic acids (cycloalkane and/or benzene
carboxylic acids with one or more saturated 5- or 6-ring carbon or aromatic
structures (Barman Skaare et al., 2007; Grewer et al., 2010). Naphthenic acids are
slightly water-soluble and, when abundant in the crude oil, also are present in the
associated produced water. Heavy crude oils, bitumens, and associated oil sands
process affected water from the Alberta, Canada, oil sands area contain high
concentrations of hundreds of naphthenic acids with eight to 30 carbons. Several
process waters from Syncrude and Suncor contain 24 to 68 mg/L total naphthenic
acids (Holowenko et al., 2002).
Produced water from the Troll C platform on the Norwegian continental self
contains highly variable concentrations and compositions of naphthenic acids,
representing different degrees of anaerobic biodegradation of crude oil in different
parts of the reservoir (Barman Skaare et al., 2007). The most abundant napththenic
acids in Troll produced water included a series of alkylated benzoic acids, salicylic
acid (2-hydroxybenzoic acid), and a variety of naphthoic acids and their ring-
reduced analogues. These organic acids were produced by anaerobic biodegradation
of aromatic hydrocarbons in the crude oil in the reservoir. Anaerobic bacteria may
be abundant in the oil/gas reservoir if formation temperature is below about 100ºC
(Yeung et al., this volume). Naphthenic acids in crude oil and produced water are of
concern because their acidity contributes to corrosion of production pipe and they
contribute to the toxicity of produced water (Thomas et al., 2009).
2.4 Petroleum Hydrocarbons
Petroleum hydrocarbons, organic chemicals consisting of just carbon and hydrogen,
are the chemicals of greatest environmental concern in produced water. Petroleum
hydrocarbons are classified into two groups: saturated hydrocarbons and aromatic
hydrocarbons. The solubility of petroleum hydrocarbons in water decreases as their
size (molecular weight) increases, aromatic hydrocarbons are more water-soluble
than saturated hydrocarbons of the same molecular weight. The hydrocarbons in
produced water appear in both dissolved and dispersed (oil droplets) form.
Existing oil/water separators, such as hydrocyclones, are quite efficient in removing
oil droplets, but not dissolved hydrocarbons, organic acids, phenols, and metals from
produced water. Thus, much of the petroleum hydrocarbons discharged to the ocean
7
in properly treated produced water are dissolved low molecular weight aromatic
hydrocarbons and smaller amounts of saturated hydrocarbons. Because there are no
treatment procedures that are 100% effective, treated produced water still contains
some dispersed oil (droplet size ranging from 1 to 10 μm) (Johnsen et al., 2004).
The droplets contain most of the higher molecular weight, less soluble saturated and
aromatic hydrocarbons (Faksness et al., 2004).
2.4.1 BTEX and Benzenes
The most abundant hydrocarbons in produced water are the one-ring aromatic
hydrocarbons, benzene, toluene, ethylbenzene, and xylenes (BTEX) and low
molecular weight saturated hydrocarbons. BTEX may be present in untreated
produced water from different sources at concentrations as high as 600 mg/L (Table
3). Produced water also contains small amounts of C3- and C4-benzenes (Table 6).
Benzene usually is most abundant and concentration decreases with increasing
alkylation (Dórea et al., 2007; Gabardo et al., this volume; Neff et al., this volume).
Because BTEX are extremely volatile, they are lost rapidly during produced water
treatment by air stripping and during initial mixing of the produced water plume in
the ocean (Terrens and Tait, 1996).
Table 6 Concentrations (mg/L) of BTEX and selected C3- and C4-benzenes in
produced water from four platforms in the U.S. Gulf of Mexico and from three
offshore production facilities in Indonesia (from Neff, 2002)
Compound 7 Gulf of Mexico
Produced Waters
3 Indonesian
Produced Waters
Benzene 0.44 — 2.80 0.084 — 2.30
Toluene 0.34 — 1.70 0.089 — 0.80
Ethylbenzene 0.026 — 0.11 0.026 — 0.056
Xylenes (3 isomers) 0.16 — 0.72 0.013 — 0.48
Total BTEX 0.96 — 5.33 0.33 — 3.64
Propylbenzenes (2 isomers) NA ND — 0.01
Methylethylbenzenes (3 isomers) NA 0.031 — 0.051
Trimethylbenzenes (3 isomers) NA 0.056 — 0.10
Total C3-Benzenes 0.012 — 0.30 0.066 — 0.16
Methylpropylbenzenes (5 isomers) NA ND — 0.006
Diethylbenzenes (3 isomers) NA ND
Dimethylethylbenzenes (6 isomers) NA ND — 0.033
Total C4-Benzenes ND — 0.12 ND — 0.068
NA: Not analyzed. ND: Not detected.
Saturated hydrocarbons, because or their low solubilities, nearly always are present
at low concentrations in produced water (Table 3), unless the produced water
treatment system is not working properly. Produced water from the U.S. Gulf of
Mexico and offshore Thailand contained 0.6 to 7.8 mg/L total C10- through C34-n-
alkanes (Neff, 2002). The shorter chain-length alkanes, C10 through C22, were more
8
abundant than the longer chain-length alkanes. Most of the alkanes probably were
associated with droplets.
2.4.2 Polycyclic aromatic hydrocarbons
Polycyclic aromatic hydrocarbons (PAH) are defined as hydrocarbons containing
two or more fused aromatic rings. These are the petroleum hydrocarbons of greatest
environmental concern in produced water because of their toxicity and persistence in
the marine environment (Neff, 1987, 2002). Concentrations of total PAH in
produced water typically range from about 0.040 to 3 mg/L (Tables 3 and 7) and
consist primarily of the most water-soluble congeners, the 2- and 3-ring PAH, such
as naphthalene, phenanthrene, and their alkyl homologues (Table 7, Figure 1).
Higher molecular weight, 4- through 6-ring PAH rarely are detected in properly-
treated produced water. Because of their low aqueous solubilities, they are
associated primarily with dispersed oil droplets (Faksness et al., 2004; Johnsen et al.,
2004). Burns and Codi (1999) reported that 5 to 10% of the total PAH in produced
water from the Harriet A platform on the northwest shelf of Australia were in the
“dissolved” fraction. The dissolved fraction contained mainly alkylnaphthalenes and
traces of alkylphenanthrenes. The particulate (droplet) fraction also contained high
concentrations of naphthalenes and phenanthrenes, and contained almost all the
dibenzothiophenes, fluoranthenes/pyrenes, and chrysenes in the produced water.
Table 7 Concentrations (μg/L = parts per billion: ppb) of individual polycyclic
aromatic hydrocarbons (PAH) or alkyl congener groups in produced water from the
Scotian Shelf and Grand Banks, Canada, the U.S. Gulf of Mexico, and the North Sea
Compound Gulf of Mexicoa North Seaa Scotian
Shelfb
Grand
Banksc
Naphthalene 5.3 - 90.2 237 - 394 1512 131
C1-Naphthalenes 4.2 - 73.2 123 - 354 499 186
C2-Naphthalenes 4.4 - 88.2 26.1 - 260 92 163
C3-Naphthalenes 2.8 - 82.6 19.3 - 81.3 17 97.2
C4-Naphthalenes 1.0 - 52.4 1.1 – 75.7 3.0 54.1
Acenaphthylene ND - 1.1 ND 1.3 2.3
Acenaphthene ND - 0.10 0.37 – 4.1 ND ND
Biphenyl 0.36 – 10.6 12.1 - 51.7 ND ND
Fluorene 0.06 - 2.8 2.6 – 21.7 13 16.5
C1-Fluorenes 0.09 - 8.7 1.1 – 27.3 3 23.7
C2-Fluorenes 0.20 – 15.5 0.54 – 33.2 0.35 4.8
C3-Fluorenes 0.27 – 17.6 0.30 – 25.5 ND ND
Anthracene ND - 0.45 ND 0.26 ND
Phenanthrene 0.11 - 8.8 1.3 – 32.0 4.0 29.3
C1-Phenanthrenes 0.24 – 25.1 0.86 – 51.9 1.30 45.0
C2-Phenanthrenes 0.25 – 31.2 0.41 – 51.8 0.55 37.1
C3-Phenanthrenes ND - 22.5 0.20 – 34.3 0.37 24.4
9
C4-Phenanthrenes ND - 11.3 0.50 – 27.2 ND 13.2
Fluoranthene ND - 0.12 0.01 – 1.1 0.39 0.51
Pyrene 0.01 - 0.29 0.03 – 1.9 0.36 0.94
C1-Fluoranthenes/
Pyrenes ND - 2.4 0.07 – 10.3 0.43 5.8
C2-Fluoranthenes/
Pyrenes ND - 4.4 0.21 – 11.6 ND 9.1
Benz(a)anthracene ND - 0.20 0.01 – 0.74 0.32 0.60
Chrysene ND - 0.85 0.02 – 2.4 ND 3.6
C1-Chrysenes ND - 2.4 0.06 – 4.4 ND 6.3
C2-Chrysenes ND - 3.5 1.3 – 5.9 ND 18.8
C3-Chrysenes ND - 3.3 0.68 – 3.5 ND 6.7
C4-Chrysenes ND - 2.6 ND ND 4.2
Benzo(b)fluoranthene ND - 0.03 0.01 – 0.54 ND 0.61
Benzo(k)fluoranthene ND - 0.07 0.006 – 0.15 ND ND
Benzo(e)pyrene ND - 0.10 0.01 – 0.82 ND 0.83
Benzo(a)pyrene ND- 0.09 0.01 – 0.41 ND 0.38
Perylene 0.04 - 2.0 0.005 – 0.11 ND ND
Indeno(1,2,3-cd)pyrene ND - 0.01 0.022 – 0.23 ND ND
Dibenz(a,h)anthracene ND - 0.02 0.012 – 0.10 ND 0.21
Benzo(ghi)perylene ND - 0.03 0.01 – 0.28 ND 0.17
Total PAHs 40 – 600 419 - 1559 2148 845
ND: Not detected
aNeff (2002).
bThebaud (DFO-COOGER Unpublished Data)
cHibernia (DFO-COOGER Unpublished Data)
2.4.3 Phenols
Concentrations of total phenols in produced water usually are less than 20 mg/L
(Table 3). Measured concentrations of total phenols in produced waters from the
Louisiana Gulf coast and the Norwegian Sector of the North Sea range from 2.1 to
4.5 mg/L and 0.36 to 16.8 mg/L, respectively (Neff, 2002; Johnsen et al., 2004). The
most abundant phenols in these produced waters are phenol, methylphenols, and
dimethylphenols. The abundance of alkyl phenols usually decreases logarithmically
with increasing number of alkyl carbons (Boitsov et al., 2007; Figure 2). Long-chain
alkylphenols with seven to nine alkyl carbons are the most toxic phenols, exhibiting
strong endocrine disruption. They are quite rare in produced water from the
Norwegian continental shelf (Figure 2). The concentration of 4-n-nonylphenol (the
most toxic alkylphenol) in produced waters from six Norwegian platforms ranged
from 0.001 to 0.012 mg/L. Five other samples did not contain detectable
concentrations of nonylphenol. The concentrations of C6- through C9-alkylphenols
are highly correlated with the concentration of dispersed oil droplets in produced
water (Faksness et al., 2004).
10
Alkylphenol ethoxylate surfactants (APE), containing octylphenols and
nonylphenols, are sometimes used in the production system to facilitate the pumping
of viscous or waxy crude oils. If the surfactant degrades, some alkylphenols may
dissolve in the produced water. Because of the toxicity of the more highly alkylated
phenols as endocrine disruptors, alkylphenol ethoxylate surfactants have been
replaced in applications where the surfactant or its degradation products may reach
the environment in significant amounts (Getliff and James, 1996).
2.5 Metals
Produced water may contain several metals in dissolved or microparticulate forms.
The type, concentration, and chemical species of metals in produced waters from
different sources is variable, depending on the age and geology of the formations
from which the oil and gas are produced (Collins, 1975) and the amount and
inorganic chemical composition of water flood water injected into the hydrocarbon
reservoir. A few metals may be present in produced waters from different sources at
concentrations substantially higher (1,000-fold or more) than their concentrations in
clean natural seawater. The metals most frequently present in produced water at
elevated concentrations, relative to those in seawater, include barium, iron,
manganese, mercury, and zinc (Neff et al., 1987; Table 8). Usually, only a few of
these metals are present at elevated concentrations in a particular produced water
sample. Concentrations of dissolved and particulate barium, iron, and manganese in
produced water from the Hibernia facility off Newfoundland are substantially higher
than concentrations in clean seawater (Yeats et al., this volume).
Injecting seawater into a well during production may increase the concentration of
sulfate (present in seawater at a concentration of about 2712 mg/L) in the formation
water, causing barium to precipitate as barite (BaSO4), and lowering the
concentration of dissolved barium in the produced water (Stephenson et al., 1994).
Several other metals in produced water, particularly radium isotopes, may co-
precipitate with barium, reducing their concentrations in produced water.
11
Table 8 Concentration ranges (μg/L = ppb) of several metals in seawater and in
produced water from the Scotian Shelf and the Grand Banks, Canada, compared to
produced water discharged to northwestern Gulf of Mexico and the Norwegian
sector of the North Sea
Metal Seawater
Gulf of
Mexicoa North Seab Scotian
Shelfc
Grand
Banksd
Arsenic 1 – 3 0.5 – 31 0.96 – 1.0 90 <10
Barium 3 – 34 81,000–
342,000
107,000–
228,000 13,500 301 – 354
Cadmium 0.001 – 0.1 <0.05 –
1.0 0.45 – 1.0 <10 <0.02 –
0.04
Chromium 0.1 – 0.55 <0.1 – 1.4 5 – 34 <1 – 10 <1
Copper 0.03 – 0.35 <0.2 12 – 60 137 <5
Iron 0.008 – 2.0 10,000–
37,000
4,200–
11,300
12,000–
28,000
1,910–
3,440
Lead 0.001 – 0.1 <0.1 – 28 0.4 – 10.2 <0.1 – 45 0.09 –
0.62
Manganese 0.03 – 1.0 1,000–
7,000 NA 1,300–
2,300 81 – 565
Mercury 0.00007 –
0.006
<0.01 –
0.2
0.017 –
2.74 <10 NA
Molybdenum 8 – 13 0.3 – 2.2 NA NA <1
Nickel 0.1 – 1.0 <1.0 – 7.0 22 – 176 <0.1 –
420 1.7 – 18
Vanadium 1.9 <1.2 NA NA <0.1 – 0.6
Zinc 0.006 –
0.12 10 – 3,600 10 – 340 10 –
26,000 <1 – 27
aCombined results from seven platforms (Neff 2002)
bCombined results from 12 platforms (Neff 2002)
cSOEP/DFO
dCombined results from Hibernia and Terra Nova (DFO-COOGER Unpublished Data)
Because formation water is anoxic, iron and manganese may be present in solution
at high concentrations. When formation waters containing high concentrations of
these metals are brought to the surface and exposed to the atmosphere, the iron and
manganese precipitate as iron and manganese oxyhydroxides. Several other metals
in produced water may co-precipitate with iron and manganese and be dispersed,
adsorbed to, or complexed with very fine, solid hydrous Fe and Mg oxides in the
receiving waters (Lee et al., 2005; Azetsu-Scott et al., 2007). Zinc and possibly lead
in produced water could be derived in part from galvanized steel structures in
contact with the produced water or with other waste streams that may be treated in
the oil/water separator system.
12
2.6 Radioisotopes
Naturally occurring radioactive material (NORM) is present in produced water in
many parts of the world. The most abundant NORM radionuclides in produced
water are the natural radioactive elements, radium-226 and radium-228 (226Ra and
228Ra). Radium is derived from the radioactive decay of uranium-238 and thorium-
232 associated with certain rocks and clays in the hydrocarbon reservoir (Reid,
1983; Kraemer and Reid, 1984; Michel, 1990). 226Ra (half-life 1,601 years) is an α-
emitting daughter of uranium-238 and uranium-234, 228Ra (half-life 5.7 years) is a β-
emitting daughter of thorium-232.
The concentration of radionuclides, such as radium, in environmental media is
measured as the rate of radioactive decay (number of disintegrations per minute),
usually as picocuries/L (pCi/L) or becquerels/L (Bq/L). A pCi is equivalent to 2.22
disintegrations per minute (dpm) or 0.037 becquerels (Bq). One pCi is equivalent to
1 pg of 226Ra or 0.037 pg of 228Ra (Neff, 2002).
The surface waters of the ocean have 226Ra and 228Ra activities of 0.027 to 0.04
pCi/L (0.027 to 0.04 pg/L: parts/quadrillion) and 0.005 to 0.012 pCi/L (0.000018 to
0.00004 pg/L), respectively (Santschi and Honeyman, 1989; Nozaki, 1991; Table 9).
Produced water from some on-shore and offshore production facilities worldwide
contain very high 226Ra and 228Ra activity (Jonkers et al., 1997), relative to activity
in nearshore and ocean waters. Concentrations of total 226Ra plus 228Ra in produced
water from oil, gas, and geothermal wells along the U.S. Gulf of Mexico coast range
from less than 0.2 pCi/L to 13,808 pCi/L (Kraemer and Reid, 1984; Fisher, 1987;
Neff et al., 1989; Hart et al., 1995; Table 9). There is no correlation between the
concentrations of the two radium isotopes in produced water from the Gulf of
Mexico, because of their different origins in the geologic formation. However, there
is a good correlation between 228Ra and 226Ra activity in produced water from the
Norwegian continental shelf, probably because most Norwegian produced water
contains very low radium activity, with a relatively few, mostly from the Sleipner
and Njord fields, containing relatively high 228Ra and 226Ra activity (NRPA, 2004).
Table 9 Mean or range of activities of 226Ra and 228Ra in produced water from
different locations, activities are pCi/L (1 pCi = 0.037 Bq)
Location Radium-226 Radium-228 Reference
Ocean water
(background) 0.027 0.04 0.005 – 0.012
Santschi &
Honeyman, 1989;
Nozaki, 1991
World-wide 0.05 – 32,400 8.1 - 4860 Jonkers et al., 1997
Scotian Shelf 1.2 9.2 Nelson, 2007
Grand Banks 33.0 229.7 Nelson, 2007
Texas 0.1 5,150 NA Fisher, 1987
13
Louisiana Gulf Coast ND 1,565 ND 1,509 Kraemer & Reid,
1984
Offshore U.S. Gulf of
Mexico 91.2 1,494 162 600 Hart et al., 1995
Santa Barbara Channel,
CA 165 137 Neff, 1997
Cook Inlet, AK <0.4 9.7 NA Neff, 1991
North Sea 44.8 105 Stephenson et al.,
1994
Dutch North Sea <54 – 8154 <27 – 540 NRPA, 2004
Norwegian continental
shelf ND – 432 ND – 567 NRPA, 2004
S. Java Sea, Indonesia 7.6 56.5 0.6 17.7 Neff & Foster, 1997
Offshore Brazil 0.5 – 294 <2 – 183 Gabardo et al., this
volume
Radium activity in produced water from offshore production areas other than the
northern Gulf of Mexico often is low (Table 9), with mean 226Ra and 228Ra activity
usually less than 200 pCi/L. Produced water from some production facilities in the
North Sea, particularly the Dutch sector, contain high 226Ra activity. Preliminary
studies of produced water from platforms discharging to Atlantic Canada show low
radium isotope activity, still several orders of magnitude higher than activity in
natural seawater. However, due to high natural dispersion, only background activity
can be detected in seawater samples collected near production platforms located on
the Grand Banks and the Scotian Shelf (Nelson, 2007). Slightly elevated radium
isotope activity has been detected near production facilities on the Norwegian
continental shelf (NRPA, 2004).
Several other radionuclides may be present at low activity in produced water. The
only radioisotope that sometimes is present in produced water at a higher activity
than that of 226Ra and 228Ra is 210Pb, a daughter of 226Ra (Table 10). Mean 210Pb
activity in treated produced water from four platforms off Louisiana ranged from
5.60 ± 5.50 pCi/L to 12.50 ± 2.60 pCi/L (Hart et al., 1995). Another daughter of
226Ra, 210Po, is present at low activity in produced water from the North Sea,
probably because of its short half-life, 138.4 days (NRPA, 2004). Activity of parents
of the radium isotopes, 238U and 232Th, are low, as is the short-lived 224Ra (half-life,
3.66 days).
Table 10 Activity (pCi/L except where noted) of radium isotopes and a few of their
parents and daughters in produced water and seawater (from Jonkers et al., 1987 and
NRPA, 2004)
Radionuclide Seawater Produced Water
226Ra 0.027 0.04 0.054 – 32,400
228Ra 0.005 – 0.03 8.1 - 4860
224Ra 0.0002 – 0.008a 13.5 - 1080
14
238U 1.1 0.008 – 2.7
232Th 0.003 0.008 – 0.027
210Pb 0.026 – 0.12b 1.35 - 5130
210Po 0.018 – 0.068b 0.005 – 0.17
aRasmussen, 2003
bCherry et al., 1987
2.7 Production Chemicals
Large numbers of specialty additives (treatment chemicals) are available for use in
the production system of a well to aid in recovery and pumping of hydrocarbons, to
protect the production system from corrosion, to facilitate the separation of oil, gas,
and water, and prevent methane hydrate (ice) formation in gas production systems.
These include biocides, scale inhibitors, emulsion-breakers, and gas-treating
chemicals (Table 11). Many of these chemicals are more soluble in oil than in
produced water and remain in the oil phase. Others are water-soluble, remain in the
produced water, and are disposed with it (Tables 11 and 12). Concentrations of most
production chemicals are low in treated produced water (Table 13). Corrosion
inhibitors, scale inhibitors, and gas treatment chemicals (glycol and methanol) may
be high in production systems with these problems.
The point in the production stream where the chemical is added influences the
amount that may be discharged or re-injected with the produced water. Treatment
chemicals are used to solve specific problems and are not added if there is no
demonstrated need. Environmental problems may arise if the more toxic treatment
chemicals, such as biocide or corrosion inhibitor, are used at a frequency or
concentration greater than needed to solve the problem. Environmental concerns
associated with the use of treatment chemicals are effectively managed through the
use of best management practices such as the Offshore Chemical Selection System
(OCSS), regulatory compliance effluent toxicity testing protocols, or use of the
DREAM model to estimate environmental impact factors (EIF) for individual
produced water treatment chemicals (Johnsen and Frost, this volume; Reed and Rye,
this volume). Production chemicals with a high EIF can be replaced with less toxic
alternatives or managed in such as way as to reduce amounts discharged to the
ocean.
Table 11 Production chemicals used on North Sea oil and gas platforms and the
estimated amounts discharged in produced water to the ocean (from Johnsen et al.,
2004)
Chemical
Typical Use
Concentration
(ppm, v/v)
Phase
Association of
Chemical
Amount
Discharged to
North Sea (t/y)
Scale inhibitor 3 – 10 Water 1143
Corrosion inhibitor 25 – 100 Oil 216
15
H2O/O2 scavenger 5 – 15 Water 22
Biocide 10 – 200 Water 81
Emulsion Breaker 10 – 200 Oil 9
Coagulants & flocculants <3 Water 197
Gas treatment chemicals Variable Water 2846
Table 12 Concentrations (mg/L) of several production treatment chemicals in
produced water discharged in the Gullfaks and Statfjord fields in the North Sea with
trade names of commercial formulations included (from Karman et al., 1995)
Chemical Concentration Chemical Concentration
Process biocide (MB554) 1.2 Well treatment
scale inhibitor
(S432)
6.8
Water injection biocide
(Kathon OM)
<0.0001 Antifoam (AF
119)
0.1
Corrosion Inhibitor (PK
6050)
1.5 Flocculent (ML
2317W)
2.4
Corrosion inhibitor (VN
6000K)
0.3 Glycol 7.7
Process scale inhibitor (SP
250)
2.1 Methanol 0.3
Process scale inhibitor (SP
2945)
0.2
3 Produced Water Treatment
3.1 Regulation of Produced Water Discharge
Most environmental regulatory agencies in countries that have significant offshore
oil and gas production place limits on the concentrations of petroleum (usually
measured as total oil and grease) that can be present in produced water destined for
ocean discharge. Table 13 summarizes examples of limits imposed by different
countries on total oil and grease or total petroleum hydrocarbon concentration in
produced water destined for ocean discharge. Different countries have proposed
different standard methods for measuring oil in produced water (Yang, this volume).
The different methods measure different fractions of the total organic chemicals in
produced water and, therefore, give different results (Veil, this volume; Yang, this
volume). For example, in the U.S., total oil and grease is defined as those materials
that are extracted by n-hexane, not evaporated at 70oC, and capable of being
weighed (gravimetric) or quantified by infrared analysis (IR). In the OSPAR
countries, total (dispersed) oil is defined as the sum of the concentrations of
16
compounds extractable with n-pentane, not adsorbed on Florisil, that can be
quantified by gas chromatography/flame ionization detection (GC/FID) with
retention times between those of n-heptane (C7H16) and n-tetracontane (C40H82),
excluding toluene, ethylbenzene, and xylenes. Gravimetric and IR methods tend to
measure many nonpolar organic chemicals in addition to petroleum hydrocarbons,
whereas GC/FID, if properly performed, measures mainly semi-volatile aliphatic
and aromatic hydrocarbons. None of the methods quantitatively measure the low
molecular weight, volatile aromatic hydrocarbons, such as BTEX and naphthalene,
that contribute to the aquatic toxicity of produced water.
Table 13 Monthly average and daily maximum concentrations (mg/L) of total oil
and grease permitted by several countries for produced water destined for ocean
disposal (from Veil, 2006)
Country Monthly Average Daily Maximum
Canada 30 60
USA 29 42
OSPAR (NE Atlantic) 30 ---
Mediterranean Sea 40 100
Western Australia 30 50
Nigeria 40 72
Brazil --- 20
Current regulatory guidelines for produced water discharge in Canada are based on
total petroleum hydrocarbon concentration, measured by IR. Under the 2002
Offshore Waste Treatment Guidelines, the hydrocarbon concentration of produced
water must be reduced to acceptable levels prior to discharge into the ocean (NEB,
2002). The minimum regulatory standard for the treatment and/or disposal of wastes
associated with the routine operations of drilling and production installations
offshore Canada is a 30-day weighted average of oil in discharged produced water of
30 mg/L and a 24-hour arithmetic mean of oil in produced water not to exceed 60
mg/L. Similar limits on total petroleum hydrocarbon concentrations in produced
water destined for ocean disposal have been promulgated by environmental
regulatory agencies in most other countries with offshore oil and gas production
(Table 13).
3.2 Produced Water Treatment
Produced water intended for ocean disposal usually is treated on the platform or at a
shore treatment facility to meet these regulatory limits. The objectives of oil-water-
gas treatment on an offshore platform are to produce stabilized crude oil and gas for
pipeline or tanker transport to shore facilities, and to generate a produced water that
meets discharge requirements (if discharged to the ocean) or is suitable for
reinjection into the producing formation or another geologic formation (Bothamley,
2004). The discharge requirements for ocean disposal are based on best available
17
technology available for treating the produced water. Research is being performed in
several countries to determine if the regulatory limits are sufficiently protective of
the marine environment.
It is necessary to treat produced water before ocean discharge to avoid the harmful
effects that the chemicals in the waste waters may have on the receiving
environment. Treatment removes solids and dispersed non-aqueous liquids from the
wastewater, including dispersed oil, suspended solids, scales, and bacterial particles,
as well as most volatile hydrocarbons and corrosive gases, such as CO2 and H2S.
Experience by the offshore oil industry with produced water treatment for ocean
disposal has shown that, if dispersed oil is removed, concentrations of volatile and
dissolved hydrocarbons are reduced to acceptable levels (Ayers and Parker, 2001). If
the treated wastewater is intended for disposal to freshwater, recycling for steam
generation for the various thermal enhanced oil recovery (EOR) technologies, or for
reinjection into the formation, most of the dissolved salts and metals also should be
removed. Salt removal is not necessary if the discharge is to the ocean.
The oil/gas/water mixture may be processed through separation devices to separate
the three phases from each other. The types of equipment used on many platforms
to remove oil and grease from produced water include mechanical and hydraulic gas
floatation units, skimmers, coalescers, hydrocyclones, and filters (Otto and Arnold,
1996; Veil, this volume). Higher molecular weight, more toxic PAH, alkylphenols,
and naphthenic acids are associated almost exclusively with dispersed oil droplets in
produced water (Fakness et al., 2004). Efficiency of removal of these toxic
chemicals can be improved by removing droplets with high-speed centrifuges and
membrane filters, capable of removing particles in the 0.01 to 2 µm range (Veil, this
volume). Chemicals may be added to the process stream to improve the efficiency of
oil/gas/water separation. The combination of mechanical and chemical treatment is
effective in removing volatile compounds and dispersed oil from the produced
water, but is ineffective in removing dissolved organics, ions, and metals. However,
even with the most advanced separation equipment, the oil/water separation is not
100% efficient.
Produced water from offshore oil and gas wells is treated to remove volatile
hydrocarbons, dispersed petroleum, and suspended solids to the extent afforded by
current waste water treatment technology. The worldwide average concentration of
total petroleum hydrocarbons in produced water discharged offshore in 2003 was 21
mg/ L, with a range for different geographic regions from 14 to 39 mg/L (OGP,
2004; Figure 3). The quality of the produced water discharged is a function
primarily of the efficiency of the treatment technologies and the strictness and
degree of enforcement of environmental discharge regulations.
Within Canada, the formulation of regulatory guidelines for produced water
discharge is an adaptive process that promotes the development of improved
environmental effects monitoring (EEM) programs and takes into consideration the
level of environmental risk, the Best Available Technology (BAT) for mitigative
measures, and the social-economic benefits.
18
4 Fate of Produced Water Following Discharge into the Ocean
4.1 Plume Dispersion Models
Treated produced water on offshore platforms may be discharged above or below the
sea surface if regulatory compliance concentrations are achieved. The location of
subsurface discharge pipes may range in depth from 10 m to 100 m. Saline
produced waters usually are as dense as or denser than seawater and disperse below
the sea surface, diluting rapidly upon discharge into well-mixed marine waters. Low
salinity produced water may form a plume on the sea surface and dilute more
slowly, particularly if discharged at or above the sea surface (Nedwed et al., 2004).
Dispersion modeling studies of the fate of produced water differ in specific details,
but all predict a rapid initial dilution of discharges by 30- to 100-fold within the first
few tens of meters of the outfall. This is followed by a slower rate of dilution at
greater distances (Terrens and Tait, 1993; Brandsma and Smith, 1995; Strømgren et
al., 1995; Smith et al., 2004). When discharge volumes of buoyant or neutral-
density produced water are very high, dilution may be slower. Factors that affect the
rate of dilution of produced water include discharge rate and height above or below
the sea surface, ambient current speed, turbulent mixing regime, water column
stratification, water depth, and difference in density (as determined by temperature
and total dissolved solids concentration) and chemical composition between the
produced water and ambient seawater (Niu et al., this volume; Reed and Rye, this
volume).
Brandsma and Smith (1995) modeled the fate of produced water discharged under
typical Gulf of Mexico conditions. They used two discharge rates: 115.7 m3/day,
which is the median flow rate for offshore discharges to the Gulf of Mexico, and
3,975 m3/day, which is the maximum allowable discharge rate from a single
discharge pipe to the Gulf of Mexico under the general National Pollutant Discharge
Elimination System (NPDES) permit. The effluent was a hypersaline brine that was
discharged at a temperature of about 29ºC. Therefore, it was denser than the
ambient seawater and tended to sink. For a median produced water discharge rate of
115.7 m3/day, the predicted concentration of produced water in the plume 100 m
down-current from the discharge ranged from 0.043 to 0.097%, depending on
ambient current speed (Table 14). At the higher discharge rate, dilutions at 100 m
down-current from the discharge ranged from 0.18 to 0.32% produced water,
depending on current speed. High-volume discharges into high current speeds in the
North Sea (10,000 m3/day) or Bass Strait off southeast Australia (14,000 m3/day)
were diluted to 1.3% and 0.47%, respectively at 100 m from the outfall (Table 14).
Brandsma (2001) estimated a rapid decrease of plume centerline concentration to
~2% at 100 m from the high-volume (6359 m3/day) the discharge from platform
Irene off Santa Barbara, California.
19
Table 14 Predicted dilutions of produced water in the receiving waters at different
discharge rates and current speeds (from Brandsma and Smith, 1995)
Location Discharge Rate
(m3/day)
Current Speed
(cm/sec)
Concentration at
100 m (% PW)
Gulf of Mexico 115.8 3.3 0.043
Gulf of Mexico 115.8 9.5 0.073
Gulf of Mexico 115.8 25.3 0.097
Gulf of Mexico 3977.8 3.3 0.18
Gulf of Mexico 3977.8 9.5 0.32
Gulf of Mexico 3977.8 25.3 0.32
North Sea 10,000 22 1.3
Bass Strait 14,000 26 0.47
High dilution rates for produced water discharges to well-mixed receiving waters
appear to be the norm. Terrens and Tait (1994) used the Offshore Operators
Committee (OOC) model to estimate dilution of a 14,000 m3 per day discharge from
the Halibut platform in Bass Strait, Australia. The produced water was dynamically
indistinguishable from the receiving waters due to the high degree of initial dilution
(dilution range of 100:1 to 252:1). At 6 km, the dilution factors were about 13,000:1
for suspended oil droplets and 18,000:1 for dissolved oil. Skåtun (1996) used a
BJET model to study the near-field mixing of a warm (T = 32°C), high salinity
(84‰) produced water released from a platform in the Gulf of Mexico. The plume
had a dilution factor of 400:1 when the current speed was 15 cm/s at 103 m
downstream from the release point, The physical dispersion models of projected
discharges from the Sable Offshore Energy Project (SOEP) wells (SOEP 1996),
located on the Scotian Shelf of Canada, also indicated that rapid dilution of the
discharge to non-acutely toxic levels within very short distances from discharge
points.
The OOC model was designed to estimate only short-term fate and other models
described above only considered the near field mixing. However, it also is important
to study the long-term, far-field transport of produced water. Hodgins and Hodgins
(1998) studied the dispersion of produced water from the Terra Nova FPSO (floating
production, storage, and offloading facility) off the east coast of Canada using the
UM3 (Three-dimensional Updated Merge) model coupled with a particle tracking-
based far-field model. For the maximum discharge of 18,000 m3/day, the estimated
worst-case initial dilution was 5:1. As the pooled effluent near the hull of the FPSO
was carried away by the ambient currents, the far field model predicted a minimum
secondary dilution of 5:1 and this yielded a combined total dilution of 25:1 after the
plume had been dispersed a few hundred meters from the FPSO. The same
modeling concept was applied to the White Rose development off the east coast of
Canada (Hodgins and Hodgins, 2000). The near-field UM3 model estimated an
initial bulk dilution of 35:1 for a discharge of produced water with a density of 728
kg per m3 at a maximum rate of 30,000 m3/day from a 36-cm pipe at 5 m below sea
surface, (Hodgins and Hodgins, 2000). The far-field dispersion simulation showed
that the 1% impact line (concentration greater than 0.1 mg L-1 or dilution < 400:1 for
20
at least 1% of the time) extended to 1.8 to 3.2 km. Similar results were described by
AMEC (2006) that predicted near-, intermediate- and far-field dilution rates with the
US EPA Visual Plumes model (Baumgartner et al., 1994), the USACE CDFate
model (Chase, 1994), and an advection/diffusion model. The models estimated a
dilution of 70:1 at 500 m from the discharge and 400:1 at 2 km from the discharge
for a maximum flow rate of 6400 m3/day of produced water from the Deep Panuke
facility off the coast of Nova Scotia.
The ASATM MUDMAP model (Spaulding, 1994) was run by Burns et al. (1998) to
study the dispersion of produced water from the Harriet oil field on the northwest
shelf of Australia. Based on an averaged daily produced water discharge of 8000
m3/day and a total oil concentration ranging from 5.9 to16 mg/L, the model
predicted that the plume had an oil concentration range from 0.006 mg/L near the
discharge to 0.00016 mg/L at about 8 km down current.
Zhao et al. (2008) integrated a random walk based particle tracking model with the
Princeton Ocean Model (POM), which enable the fast prediction of future dispersion
and risks of produced water discharges. For a produced water discharge of 21,200
m3/day from the Hibernia platform off the east coast of Canada, the model predicted
a Pb concentration of 0.002 μg/L (about the concentration in clean seawater: Table
8) at about 5.3 km south of the platform. An overestimation of dilution and
underestimation of chemical concentration may result from the approach of Zhao et
al. (2008) due to omission of near-field buoyant jet behaviors. The accuracy of the
model can be improved by coupling a near-field model with the particle-tracking
model using the method described by Zhang (1995) and Niu et al. (this volume).
Mukhtasor et al. (2004) estimated a mean concentration of 0.5% the initial
concentration of the produced water plume at about 225 m from the discharge in a
hypothetical study of the produced water discharge from the Terra Nova FPSO. The
95%-tile concentration from their analysis at the same location was about 2.25% the
initial concentration. Similar approaches were used by Niu et al. (2009a) to study
the effects of surface waves on the dispersion of produced water. Because the
models of Mukhtasor (2004) and Niu et al. (2009a) can only be used for a limited
number of discharge conditions, Niu et al. (2009b; this volume a, b) expanded the
same approach into a probabilistic based steady state model (PROMISE) that was
coupled to a MIKE3 model to study dispersion over a wide range of non-steady state
discharge and environmental conditions. Niu et al. (2009b; this volume a, b)
performed a validation study of the PROMISE/MIKE3 model and reported good
agreement between model predictions and laboratory measurements.
As is evident in this overview, during the last two decades, a significant amount of
effort has been put forward to model the dispersion of produced water plumes in the
marine environment. Researchers from different disciplines have approached the
problem from different perspectives and developed models with various degrees of
sophistication. Produced water plume dispersion models have evolved from the
simple steady state near-field, short term dilution models to comprehensive coupled
hydrodynamic/dispersion models that predict both the near- and far-field dispersion
21
processes in 3-D non-steady state conditions. Although the dispersion models are
now able to simulate the near-field mixing process very well and predict the far-field
mixing process reasonably well, they are still limited in their capacity to predict the
fate of the various chemical components of produced water.
4.2 Chemical Fate/Transport Models
In the majority of physical transport models, the chemicals in the produced water
stream are treated as passive tracers. Similarly, dye injection tracer studies of
produced water plumes (e.g., DeBlois et al., 2007) are based on the same
assumptions. The drawback of using dyes is that they may not become fully
integrated within the produced water plume, resulting in an inaccurate prediction of
the transport of various contaminants that react chemically and separate from the
plume. Based on recent studies by Lee et al. (2005a) and Azetsu-Scott et al. (2007),
consideration also must be given to the chemical transformations that may influence
the subsequent transport, fate, and effects of the contaminants of concern in the
produced water following its discharge.
Berry and Wells (2004a, b) first used the CORMIX model (Doneker and Jirka,
2007) to determine the exposure pathways and potential compartment interactions
and then employed a Level III fugacity model (Mackay, 1991) to study the
distribution of benzene and naphthalene among environmental media such as water,
suspended particles, fish, and sediments from produced water discharges from the
Thebaud platform on Scotian shelf. They predicted that the averaged water column
bulk concentrations of benzene and naphthalene over a 1 km by 1 km area are 5.28 ×
10-6 μg/g and 8.49 × 10-7 μg/g (~5.28 and 0.85 ng/L (parts per trillion), respectively)
for the maximum discharge rate of 211.7 m3/day. These concentrations are close to
the measured background concentrations of benzene and naphtalene in ocean waters
(~1 ng/L: Neff, 2002). Because the CORMIX model predicted that the produced
water plume may not be fully mixed within the selected compartments, the fugacity
model may underestimate the concentration within the produced water plume and
overestimate the concentration outside of the plume.
Smith et al. (1996), used a coupled model to simulate the transport of produced
water from the Pertamina/Maxus operation area in Java Sea, Indonesia. The
CORMIX model (Doneker and Jirka, 2007) was first used to predict the effluent
dispersion and the results were then used in the PISCES model (Turner et al., 1995)
to study the partitioning, degradation, and volatilization of chemicals in the
produced water plume. The model predicted that the concentration of mercury was
about 0.0055 to 522 µg/L at 500 m from the discharge and the concentration of
arsenic was about 5 to 12 µg/L. At 3000 m downstream, the mean mercury and
arsenic concentrations decreased to 65 µg/L and 3 µg/L, respectively. These
concentrations are about 10,000 times and slightly higher than the mercury and
arsenic concentrations in clean seawater (Table 8).
22
The use of separate dispersion, fate, and chemical models cannot account for the
dynamic changes of chemical concentrations following discharge. Murray-Smith et
al. (1996) applied the particle-based model, TRK, to the Clyde platform in the UK
sector of the North Sea. The model combined physical dispersion with first order
degradation to simulate biodegradation and other removal processes. The study
found that the initial dilution was rapid, with a minimum dilution factor of between
300 and 3000 within 100 m from the discharge. Further away, the physical dilution
was less rapid and other removal processes, such as precipitation, biodegradation
may have become important. The model predicted an overall dilution (including
biodegradation) of plume chemicals of 1000-16,000-fold at 1 km from the discharge.
Reed et al. (1996) described a PROVANN model for the simulation of 3-
dimensional transport, dilution, and degradation of chemicals associated with
produced water discharges from one or more simultaneous sources. The model
included various transport processes such as adsorption/dissolution kinetics,
entrainment and dissolution of oil droplets, volatilization, degradation, and
deposition from water column. For two platforms off Trondheim, Norway, the
simulation showed that the naphthalene concentration at the edge of the plume at
about 40 km downstream after 50 days was extremely low (~0.00006 µg/L). The
simulation showed that inclusion of degradation is clearly an important factor in
modeling long term fate of produced water chemicals.
PROVAN evolved over several years to the Dose-related Risk and Exposure
Assessment Model (DREAM), a 3-dimensional, time-dependent numerical model
that computes transport, exposure, dose, and effects in the marine environment
(Reed et al., 2001; Reed and Rye, this volume). Reed and Rye (this volume) describe
DREAM and its application in modeling the fate and effects of produced water
discharges to the marine environment (Figure 4).
4.3 Model Validation with Field Measurements
Field measurements are important for both understanding of the fate of produced
water and for model validation. Traditionally, this was achieved by either collecting
water samples at pre-determined stations or by continuous towing of a fluorometer
(Murry-Smith et al., 1996; Smith et al., 1994; Smith et al., 1996; Smith et al., 2004;
Terrens and Tait, 1996). Traditional ship-based sampling methods often are
expensive and time consuming and, therefore, only limited information can be
collected. The greater water depth of recent offshore production activities also
increases the level of sampling error. Recently, new and innovative means of
conducting field measurements using Autonomous Underwater Vehicles (AUV)
have been proposed and may be used in future studies of produced water fate and to
collect data to validate mathematical fate/transport models (Niu et al., 2007; Niu et
al., this volume a).
23
Field measurements of produced water dilutions usually are highly variable but
confirm the predictions of modeling studies, that dilution usually is rapid. The
comparison of field measurements of concentrations of hydrocarbons and various
other organic produced water components from both fixed sampling stations and
continuous towed fluorometers with modeled data showed that measured dilutions
were generally much higher than predicted (Murray-Smith et al., 1996). The
measured dilutions for alkanes and methylnapthalenes at 100 m to 1000 m from the
discharge ranged from 5000- to 50,000-fold, much greater than predicted. The study
also found that the measured concentrations of other organics such as phenol and
methylphenol, in the water column were below detection limits, probably because of
their rapid biodegradation. Similar results were reported for the concentrations of
five heavy metals in produced water samples discharged to the Java Sea where a
high dilution ratio was observed as predicted in the models (Smith et al., 1996).
Results of the DREAM model have been validated with field measurements in the
North Sea. Concentrations of PAH were measured in the water column and in blue
mussels (Mytilus edulis) deployed at different distances from production platforms
in the Norwegian sector off the North Sea (Johnsen et al., 1998; Røe Utvik et al.,
1999; Durell et al., 2006; Neff et al., 2006). Direct measurements of PAH in the
water gave inconsistent results, because concentrations were too low and variable.
However, the mussels did bioaccumulate PAH from the water. PAH concentrations
in mussels decreased with distance down-current from platforms. The advantage of
using mussels to estimate concentrations of nonpolar organic chemicals in the water
column is that they integrate (average) water concentrations over time, whereas
discrete water samples or fluorometer surveys can only measure concentrations at a
single time.
PAH residues in mussel tissues were used to estimate PAH concentrations in surface
waters (Neff and Burns, 1996). Estimated surface water total PAH concentrations
ranged from 0.025 to 0.35 μg/L (parts per billion) within 1 km of platform
discharges and reached background levels of 0.004-0.008 μg/L within 5-10 km of
the discharge, representing a 100,000-fold dilution of the total PAH concentration in
the discharge water (Durell et al., 2006; Neff et al., 2006). Dilution modeling
showed that most of the produced water plume was restricted to the upper 15 m to
20 m of the water column. Dilution was very rapid.
Total and individual PAH concentrations in the upper water column, estimated from
residues in mussel tissues, were compared to concentrations predicted by the
DREAM model (Durell et al., 2006; Neff et al., 2006). There was very good
agreement for measured and modeled concentrations of individual and total PAH in
the Ekofisk but not the Tampen field. The poor predictions in the Tampen field were
caused by lack of accurate physical oceanographic data for the period of the mussel
deployment. DREAM tended to overestimate naphthalenes concentrations and
underestimate concentrations of 4- and 5-ring PAH (Figure 5).
24
Harman et al. (2010) confirmed these results by deploying passive sampling devices,
semipermeable membrane devices (SPMDs) and polar organic integrative chemical
samplers (POCIS), in the vicinity of an offshore oil production platform discharging
produced water to the North Sea. There was a gradient of decreasing concentrations
in the receiving waters of low molecular weight PAH and alkylphenols with distance
from the platform. However, there was no gradient of concentrations of high
molecular weight PAH with distance from the platform.
The DREAM model predicted that the concentrations of PAH and other chemicals in
the produced water plume at different distances down-current from the discharge
exhibited wide cyclic concentration variations due to tidal and wind-driven current
flows (Figure 6). Because of rapid dilution and fluctuating water-column
concentrations, the model predicted that potentially toxic concentrations and contact
times of PAH would not occur even in the near-field.
Terrens and Tait (1996) measured concentrations of BTEX and several PAH in
ambient seawater 20 m from a 11,000-m3/day produced water discharge from a
platform in the Bass Strait off southeastern Australia (Table 15). There was an
inverse relationship between molecular weight (and volatility) and the dilution of
individual aromatic hydrocarbons. Twenty meters down-current from the produced
water discharge containing an average of 6,410 μg/L BTEX, the average BTEX
concentration in the plume was 0.43 μg/L, a dilution of 14,900-fold. PAH were
diluted by 11,000-fold (naphthalene) to 2,000-fold (pyrene). Concentrations of
higher molecular weight PAH were below the detection limit (0.0002 μg/L) in the
ambient seawater 20 m from the outfall. The inverse relationship between aromatic
hydrocarbon molecular weight and their rates of dilution probably was due in large
part to the high temperature (95ºC) of the discharged produced water, favoring
evaporation of the lighter aromatic hydrocarbons.
Table 15 Concentrations (μg/L = parts per billion: ppb) of toluene and three PAH in
the produced water discharged from the Kingfish B Platform in the Bass Strait,
Australia, and in near-surface seawater collected 20 m down-current from the
discharge (from Terrens and Tait, 1996)
Chemical Kingfish
Produced Water
Ocean 20 m from
Discharge Dilution
Toluene 3000 0.18 16,700
Naphthalene 440 0.04 11,000
Phenanthrene 18 0.003 6000
Fluoranthene 0.8 0.0002 4000
Continental Shelf Associates (1993) measured a dilution factor of 426 at 5 m and of
1,065 at 50 m from the discharge for 226Ra in a 1070 m3/day produced water
discharge plume in a water depth of 18 m in the Gulf of Mexico. The current speed
at the time of the measurements was 15 cm/s and there was little vertical
stratification of the water column. The OOC (but not the CORMIX) model
25
accurately predicted the dilutions measured in the field. Produced waters from the
Gulf of Mexico often contain high concentrations of dissolved barium. It is probable
that the radium in the produced water co-precipitated rapidly with barium sulfate in
the sulfate-rich receiving waters.
In summary, the majority of plume dispersion and chemical fate/transport models
developed to date focus on the process of dispersion and treat produced water as a
single conservative contaminant. Only a few models have attempted to include other
transformation processes, such as biodegradation, metal speciation, evaporation, and
adsorption. Among these models, the most comprehensive appears to be DREAM
that is capable of handling a multitude of complex processes and data including
discharge volumes, physical, chemical and biological fates of discharged substances,
biological uptake and effects (Reed et al., 2001; Reed and Rye, this volume).
DREAM is capable of predicting the fate of individual chemicals associated with
produced water and currently is used extensively by North Sea operators to achieve
the regional regulatory goal of “zero harmful discharges”. The model can also be
used to predict environmental effects by two approaches: the environmental impact
factor (EIF) and a body burden related risk assessment model (Johnsen and Frost,
this volume; Reed and Rye, this volume).
5 Environmental Effects of Produced Water Discharges
Based on the concentrations and relative toxicity of chemicals in most produced
waters and predicted dispersion and biodegradation/transformation rates in the
receiving waters, it is likely that there is only a limited potential for acute toxicity
beyond the immediate vicinity of produced water discharges to offshore waters of
Atlantic Canada. This hypothesis is supported by sensitive biotests – primarily
regulatory acute toxicity assays, and the rapid dispersion and degradation of the
produced water plume in the receiving waters (Lee et al., 2005a). However,
Holdway (2002) proposed that the chronic impacts associated with long-term
exposures must be quantified to fully assess the potential long-term ecological
impact of produced water discharges. Continual chronic exposure may cause sub-
lethal changes in populations and communities, including decreased community and
genetic diversity, lower reproductive success, decreased growth and fecundity,
respiratory problems, behavioral and physiological disorders, decreased
developmental success and endocrine disruption.
Fisheries and Oceans Canada and other environmental regulatory agencies
worldwide are performing ongoing chronic toxicity studies to support the
development of cost-effective and sensitive monitoring and environmental
assessment protocols for regulatory use.
26
5.1 Potential for Effects in Water-Column Organisms
Harmful biological effects in water-column biological communities near open-ocean
produced water discharges are expected to be minimal and localized, because of the
rapid dilution, dispersion, and transformation rates of most produced water
chemicals. However, some produced waters contain chemicals that are highly toxic
to sensitive marine species, even at low concentrations. When discharge is to
shallow, enclosed coastal waters, or when discharge is of a low-density produced
water in an area with low water turbulence and current speeds, concentrations of
produced water chemicals may remain high for long enough to cause ecological
harm (Neff, 2002). The chemicals of greatest environmental concern in produced
water, because their concentrations may be high enough to cause bioaccumulation
and toxicity include aromatic hydrocarbons, some alkylphenols, and a few metals.
Highly alkylated phenols (octyl- and nonyl-phenols) are well-known endocrine
disruptors, but rarely are detected in produced water at high enough concentrations
to cause harm to water column animals following initial dilution (Thomas et al.,
2004; Boitsov et al., 2007; Sundt et al., 2009). Most metals and naturally-occurring
radionuclides are present in produced water in chemically reactive dissolved forms
at concentrations similar to or only slightly higher than concentrations in seawater
and, therefore, are unlikely to cause adverse effects in the receiving water
environment (Neff, 2002). Nutrients (nitrate, phosphate, ammonia, and organic
acids) may stimulate microbial and phytoplankton growth in the receiving waters
(Rivkin et al., 2000; Khelifa et al., 2003). Some production treatment chemicals are
toxic and, if they are discharged at high concentration in produced water, could
cause localized harm (Sverdrup et al., 2002). Inorganic ions (e.g., sodium,
potassium, calcium, and chloride) are not of concern in produced water discharges to
the ocean (Pillard et al., 1996), but are of environmental concern when the treated
water is discharged to land or surface fresh or brackish waters.
5.2 Potential for Accumulation and Effects in Sediments
If produced water is discharged to shallow estuarine and marine waters, some metals
and higher molecular weight aromatic and saturated hydrocarbons may accumulate
in sediments near the produced water discharge (Neff et al., 1989; Means et al.,
1990; Rabalais et al., 1991), possibly harming bottom living biological communities.
In well-mixed estuarine and offshore waters, elevated concentrations of saturated
hydrocarbons and PAH in surficial sediments sometimes are observed out to a few
hundred meters from a high-volume produced water discharge. The concentrations
of PAH in sediments near offshore produced water discharges are related to the
volume and density of produced water discharged, the PAH concentration in it,
water depth, and local mixing regimes. PAH in sediments near offshore platforms
also may come from drilling discharges, particularly if oil based drilling muds are
used and oily drill cuttings are discharged, a practice no longer allowed in most
marine waters (Neff, 2005).
27
Barium, iron, and manganese are the metals most often greatly enriched in produced
waters compared to their concentrations in natural seawater. Speciation occurs
following the ocean discharge of produced water, in which the metals precipitate
rapidly when produced water is discharged to well-oxygenated surface waters
containing a high natural sulfate concentration. Trefry and Trocine (this volume)
showed that dissolved barium in produced water precipitates more slowly than
predicted when the produced water is discharged to sulfate-rich seawater. However,
precipitation of barium and dilution of the resulting barite in the produced water
plume are rapid enough that dissolved barium concentrations rarely exceed acutely
toxic concentrations. Other alkaline earths, such as strontium, magnesium, and
radium, co-precipitate with barium and are rapidly deposited with barium sulfate in
bottom sediments (Neff, 2002; Trefry and Trocine, this volume).
Dissolved iron and manganese precipitate rapidly as oxyhydroxides when the anoxic
produced water plume mixes with oxygen-rich receiving waters. The extremely fine-
grained iron and manganese oxides adsorb to or co-precipitate with several other
metals from the produced water plume (Lee et al., 2005a; Azetsu-Scott et al., 2007).
These particulate metals tend to settle slowly out of the water column and
accumulate to slightly elevated concentrations in surficial sediments over a large
area around the produced water discharge (Neff, 2002; Lee et al., 2005a). In
addition, the transport and concentration of inorganic constituents within produced
water (e.g., metals) to the surface microlayer may be promoted by the interaction
between residual oil droplets and metal precipitates (Burns et al., 1999; Lee et al.,
2005a). Toxicity assessment using the Microtox® test, a regulatory bioassay
protocol based on inhibition of a primary metabolic function of a bioluminescent
bacterium, showed that unfiltered produced water samples containing metal
precipitates generally had higher toxicity than filtered samples (Azetsu-Scott et al.,
2007). Current results from regulatory environmental effects monitoring programs
generally show that natural dispersion processes appear to control the concentrations
of toxic metals in the water column and sediments just slightly above natural
background concentrations.
5.3 Aquatic Toxicity of Produced Water
Most treated produced water has a low to moderate toxicity. A typical distribution of
produced water toxicities can be seen in data for produced waters discharged to the
Gulf of Mexico off the Louisiana coast (Table 16). A small number of produced
water samples are moderately toxic to mysids (a small shrimp-like crustacean) and
sheepshead minnows, with acute and chronic toxicities less than 0.1 percent (1,000
mg L-1) produced water. A few produced waters are practically nontoxic with acute
and chronic toxicities higher than 35 or 40%. Most produced waters have moderate
toxicities, with acute and chronic toxicities between about 2 and 10% for mysids and
5 to 20% for sheepshead minnows. Based on earlier toxicity studies for produced
28
waters from the Gulf of Mexico, Neff (1987) reported that nearly 52 percent of all
median lethal concentrations (LC50) were greater than 10% produced water, 37
percent were between 1 and 9.9%, and 11 percent were less than 1%. These toxicity
threshold limits are consistent with those reported for Atlantic Canada. A 1:100
dilution of the produced water, as usually occurs within a few tens of meters of the
discharge pipe, would render all but a few of these produced water samples not
acutely toxic.
Table 16 Acute and chronic toxicity of more than 400 produced water (PW) samples
from the Gulf of Mexico off Louisiana, USA, to mysids (Mysidopsis bahia) and
sheepshead minnows (Cyprinodon variegatus), exposure concentrations are percent
produced water (from Neff, 2002)
Test Number
of Tests
Mean Value
(% PW)
Standard
Deviation
Maximum
Value
Mysidopsis bahia
96h Acute Toxicity 412 10.8 10.4 86.3
Chronic Survival (NOEC) 407 3.4 5.8 50.0
Chronic Growth (NOEC) 391 2.4 3.6 42.0
Chronic Fecundity (NOEC) 274 2.7 3.2 25.0
Cyprinodon variegates
96h Acute Toxicity 359 19.2 14.8 >100
Chronic Survival (NOEC) 401 6.3 9.0 >100
Chronic Growth (NOEC) 395 5.2 8.1 >100
NOEC: no observed effect concentration
In a comprehensive study on the acute effects of produced water recovered from a
Scotian Shelf offshore well on the early life stages of haddock, lobster and sea
scallop in terms of survival, growth and fertilization success, Querbach et al. (2005)
noted that fed, stage I lobster larvae were the most sensitive with an observed LC50
of 0.9%. Feeding stage haddock larvae and scallop veligers were the least sensitive
with LC50 values of 20 and 21% respectively. In terms of chronic responses, the
average size of scallop veligers was significantly reduced after exposure to produced
water concentrations >10%.
There are poorly characterized species differences in the toxicity of produced waters
to marine organisms. When bioassays were performed with two or more marine
taxa and the same sample of produced water, crustaceans were generally more
sensitive than fish (Neff, 1987; Louisiana Department of Environmental
Conservation, 1990; Jacobs and Marquenie, 1991; Terrens and Tait, 1993). Mixed
function oxigenase (MFO) enzyme activity was highly induced in the liver, gills and
heart of juvenile cod exposed to 5% produced water for 72 hours (Andrews et al.,
2007). However, no differences in mortality were observed between control and
experimental copepods (Calanus finmarchicus), a major prey species for fish in the
29
northwest Atlantic, when they were exposed to 5% produced water for 48 hours
(Payne et al., 2001a).
Gamble et al. (1987) introduced produced water at a concentration equivalent to a
400 – 500 fold dilution of produced water (expected within 0.5 to 1.0 km from the
Auk and Forties platforms in the North Sea) into 300 m3 mesocosm tanks containing
natural assemblages of phytoplankton, zooplankton, and larval fish. Bacterial
biomass increased but phytoplankton production and larval fish survival were
unaffected in the produced water-dosed containers. However, early life stages of
copepods were sensitive to the produced water and suffered high mortalities. The
decrease in zooplankton abundance resulted in an increase in the standing stock of
phytoplankton and a reduction in the growth rates of the fish larvae. In other
mesocosm studies summarized by Stephenson et al. (1994), larval mollusks and
polychaete worms also were adversely affected. These mesocosm studies show that
low concentrations of produced water may have subtle effects on marine planktonic
communities. However, it should be pointed out that mesocosm studies represent
conservative, worst-case exposure scenarios, because produced water chemicals in
the mesocosm enclosures do not degrade and disappear as rapidly as they do in well-
mixed ocean environments.
5.4 Bioaccumulation and Biomarkers as Evidence of Exposure
Bioaccumulation is the uptake and retention of a bioavailable chemical from one or
more possible external sources such as water, food, substrate or air (Neff, 2002).
Marine animals near a produced water discharge may bioaccumulate metals,
phenols, and hydrocarbons from the ambient water, their food, or bottom sediments.
An attempt was made to measure bioaccumulation of four metals (arsenic, barium,
cadmium, and mercury), BTEX, phenol, and PAH by two species of bivalve
molluscs from platform legs and five species of fish collected within 100 m of
produced water discharging and non-discharging platforms in the Gulf of Mexico
(Neff et al., this volume). There was no difference in concentrations of any of the
metals, phenol, or BTEX in tissues of bivalves and fish from discharging and non-
discharging platforms. Concentrations of total PAH were low and highly variable in
tissues of the two bivalves and five species of fish from all platforms. Total PAH
concentrations were higher (usually by an order of magnitude or more) in the
mollusc tissues than in the fish tissues, probably because of the high activity of
PAH-metabolizing enzymes in fish. Total PAH concentrations were significantly
higher in tissues of one or both species of bivalves from the discharging platforms
than from the reference platforms at the time of one or both field surveys. Alkyl
naphthalenes, phenanthrenes, or dibenzothiophenes, all characteristic of petroleum
sources, were the individual PAH that were present most frequently at elevated
concentrations in bivalves from discharging platforms. Thus, there was evidence of
exposure to and bioaccumulation of PAH, but not metals, phenol, or BTEX from
30
produced water by bivalves associated with the biofouling community on submerged
structures on produced water discharging platforms in the Gulf of Mexico.
Biomarkers are biochemical, physiological, or histological changes in an organism
caused by exposure to and bioaccumulation of specific chemicals in water, food, or
sediments (Forbes et al., 2006). Biomarkers usually are not direct indicators of
harmful effects caused by the exposure, but can be used as early warnings of
possible risk to the exposed organism. The most useful biomarkers respond to a
single or small group of chemical contaminants and, so, can be used as evidence of
exposure to a particular class of chemicals. For example, any of several measures of
the induction (increase in activity) of the enzyme system, cytochrome P450 mixed
function oxygenase (CYP1A or MFO), can be used as evidence of exposure to PAH,
polychlorinated biphenyls (PCB), and any of several chlorinated hydrocarbon
pesticides.
As discussed above, produced water often is a source of PAH in waters and
sediments near offshore oil and gas production facilities. Exposure to and effects of
PAH, including those in produced water, have been the focus of numerous
laboratory and field studies using endpoints based on biochemical, histopathological,
immunological, genetic, reproductive, and developmental parameters (Neff, 2002;
Payne et al., 2003).
Børseth and Tollefsen (2004) monitored bioaccumulation and biomarker responses
in mussels (Mytilus edulis) and Atlantic cod (Gadus morhua) held in cages in the
vicinity of the Troll B Platform on the Norwegian continental shelf. Cages were
deployed for six weeks both inside (500 and 1,000 m from the source) and outside
the zone of expected influence of the produced water plume. They reported that
concentrations of metals and PAH in soft tissues of the caged mussels correlated
well with distance from the discharge, with highest body burdens in mussels closest
to the platform. The PAH assemblage in mussel tissues was dominated by alkyl
homologues of naphthalene, phenanthrene, and dibenzothiophene, suggesting
exposure to PAH from the produced water discharge. Biomarker responses in the
mussels were weak, providing only equivocal evidence of exposure to produced
water chemicals.
Durell et al. (2006) and Neff et al. (2006) confirmed the results of Børseth and
Tollefsen (2004) with mussels deployed at different distances from production
platforms in the Ekofisk and Tampen fields off Norway. Concentration of total PAH,
decalins (decahydronaphthalenes), and heterocycyclic aromatic compounds
(dibenzothiophenes) in mussel tissues decreased with distance from production
platforms in the Ekofisk field (Table 17). Concentrations of total PAH in the water
were estimated from tissue residues in the mussels, and were low at all distances
from the produced water discharges. The PAH assemblage in the nearfield receiving
waters (0.5 km) was dominated by alkyl-decalins, and naphthalenes. Decalins were
lost rapidly from the water column with distance from the discharges, probably
because of their relatively high volatility (Durell et al., 2006). There were only traces
of the 4- through 5-ring PAH that are largely responsible for CYP1A induction
31
(Neff, 2002). Concentrations in the water of these high molecular weight PAH did
not decrease much with distance from the discharge, probably because of deposition
of pyrogenic PAH from the atmosphere throughout the area.
Børseth and Tollefsen (2004) found no significant difference in levels of plasma
vitellogenin (an indicator of exposure to endocrine-disrupting chemicals) in male
cod from exposed and reference sites. No significant differences were detected in
ethoxyresorufin-o-deethylase (EROD) activity (biomarker of exposure to chemicals,
including PAH, that induce the cytochrome P450 mixed function oxygenase enzyme
system) in livers of fish from exposed and reference locations, indicating little or no
exposure to PAH. Levels of PAH metabolites in cod bile were low, confirming the
low level exposure to PAH. Concentrations of naphthalene metabolites in cod bile
decreased with distance from the platform, indicating that the low-level exposure to
PAH was probably from the platform‘s produced water discharge. Other biomarkers
showed little or no evidence that the cod were exposed to chemicals from the
produced water plume. The authors concluded that mussels and cod deployed near a
produced water discharge probably were exposed to low concentrations of produced
water chemicals, below levels that might represent a health risk to water-column
organisms. The low biomarker responses can be explained by the low concentrations
of PAH, particularly the higher molecular PAH that are the strongest inducers of
CYP1A biomarkers (Durell et al., 2006; Neff et al., 2006).
Table 17 Concentrations of total PAH, decalins, and dibenzothiophenes in tissues of
mussels (Mytilus edulis) following deployment at different distances down-current
from production platforms in the Ekofisk field off Norway and estimated
concentrations of total PAH in the receiving waters based on PAH residues in
mussel tissues (from Neff et al., 2006)
Distance from PW
discharge (km)
Mussels
(ng/g dry wt)
Water
(µg/L)
0.5 8630 0.086
5 2710 0.025
10 231 0.008
20 100 0.005
90 189 0.006
As part of the Biological Effects of Contaminants in Pelagic Ecosystems
(BECPELAG) Program, bioaccumulation and several biomarkers were measured in
wild and caged marine animals along a transect away from a Statfjord platform in
the North Sea (Hylland et al., 2006). Produced water discharge is 74,100 m3 per day
from three platforms in the Statfjord field (Durell et al., 2006), among the highest
discharge rates of any offshore field in the world. Førlin and Hylland (2006)
measured hepatic EROD activity and bile metabolites in juvenile cod caged at
several distances down-current from one of the discharges. There were no significant
trends in EROD activity in male and female cod with distance from the discharge,
though there was a trend for EROD activity in female cod to increase with distance
from the discharge, a trend opposite the expected one (Figure 7). However,
32
concentrations of alkyl naphthalene metabolites (alkylnaphthalenes are abundant in
produced water) in fish bile were highest in cod near the platform and decreased
with distance from the platform (Figure 8). There were no distance trends in
concentrations of other PAH metabolites in cod bile. The authors concluded that the
cod were exposed to low levels of PAH from the produced water discharges, but
exposure levels were well below those that would pose a health risk to fish living
near the platforms.
Sturve et al. (2006) exposed juvenile Atlantic cod (Gadus morhua) to North Sea oil,
nonylphenol and a combination of the North Sea oil and an alkylphenol mixture in a
flow-through system. Several hepatic biomarkers were monitored. Although
exposure to North Sea oil resulted in strong induction of CYP1A protein levels and
EROD activities, exposure to oil plus nonylphenol resulted in decreased CYP1A
levels and EROD activities. Thus, nonylphenol appeared to down-regulate CYP1A
expression in Atlantic cod. Meier et al. (2007, 2010) described the effects of
produced water and alkylphenols (AP) on early life stages and the reproductive
potential of first-time spawning Atlantic cod. Cod were fed with feed paste
containing a mixture of four reference alkylphenols, at a range of concentrations for
either 1 or 5 weeks. The AP-exposed female fish had impaired oocyte development,
reduced estrogen levels, and an estimated delay in the time of spawning of 17-28
days. Male AP-exposed fish had impaired testicular development, with an increase
in the amount of spermatogonia and a reduction in the amount of spermatozoa
present. Meier et al. (2007) concluded, based on the results of these laboratory
studies, that AP associated with a produced water discharge may have a negative
influence on the overall reproductive fitness of cod populations.
Abrahamson et al. (2008) exposed juvenile cod for two weeks in the laboratory to
several concentrations of produced water from the Oseberg C platform in the
Norwegian North Sea. There was a dose-related increase in gill EROD activity in the
fish. However, when cod were caged for six weeks at 0.5 to 10 km from the Troll B
and Statfjord B platforms, gill EROD activity was low in all fish and there was not a
clear gradient of decreasing activity with increasing distance from the discharge.
Thus, the concentration of higher molecular weight PAH in the produced water
plume usually is not high enough at 0.5 to 1.0 km from North Sea produced water
discharges to induce hepatic and gill EROD activity.
Elevated CYP1A enzyme activity has been observed in fish larvae collected
downstream of the Hibernia field (Payne et al., 2003). However, induction may be
occurring only near the platform site with the induced larvae being transported
downstream by currents. Petro-Canada and Husky Energy have performed
biomarker studies with American plaice (Hippoclossoides platessoides) collected in
the vicinity of the Terra Nova and the White Rose offshore developments on the
Grand Banks of Newfoundland (DeBlois et al., 2005; Husky Energy, 2005; Mathieu
et al., this volume). These studies showed that the overall health of the American
plaice collected in the vicinity of Terra Nova and White Rose was similar to the
health of American plaice collected at distant reference sites. American plaice are
33
demersal (bottom-living) fish and may not have been exposed to the produced water
plume in the upper water column.
Meier et al. (2010) exposed cod to produced water during the embryonic, early
larval (up to 3 months of age), or the early juvenile stages (from 3 to 6 months of
age). Alkylphenols bioconcentrated in fish tissue in a dose and developmental stage
dependent manner during produced water exposure. However, juveniles appeared
able to effectively metabolize the short chain APs. However, exposure to produced
water had no effect on embryo survival or hatching success. However, 1% produced
water, but not 0.01 or 0.1%, interfered with development of normal larval
pigmentation. After hatching, most of the larvae exposed to 1% produced water
failed to begin feeding and died of starvation. This inability to feed was linked to an
increased incidence of jaw deformities in exposed larvae. Cod exposed to 1%
produced water, had significantly higher levels of the biomarkers, vitellogenin and
CYP1A, in plasma and liver, respectively.
Hamoutene et al. (this volume) investigated the effects of produced water on cod
immunity, feeding, and general metabolism by exposing fish to diluted produced
water at concentrations of 0, 100 and 200 ppm for 76 days. No significant
differences were observed in weight gain or food intake. Similarly, serum
metabolites, whole blood fatty acid percentages, and mRNA expression of a brain
appetite-regulating factor (cocaine and amphetamine regulated transcript) remained
unchanged between groups. Other than an irritant-induced alteration in gill cells
found in treated cod, resting immunity and stress response were not affected by
produced water. Catalase and lactate dehydrogenase changes in activities were
recorded in livers but not in gills, suggesting an effect on oxidative metabolism
subsequent to hepatic detoxification processes. To evaluate potential effects of
produced water discharges on cod immunity, fish from the three groups were
challenged by injection of Aeromonas salmonicida lipopolysaccharides (LPS) at the
end of exposure. LPS injection affected respiratory burst activity of head-kidney
cells, and circulating white blood cell ratios, and increased serum cortisol in all
groups. The most pronounced changes were seen in the group exposed to the
highest dose of produced water (200 ppm).
In a followup study, Pérez-Casanva et al. (2010) reported that chronic exposure of
juvenile cod to produced water had no significant effects on growth, hepatosomatic
index, condition factor or plasma cortisol. The immune response of respiratory burst
(RB) of circulating leukocytes was significantly elevated and the RB of head–kidney
leukocytes was significantly decreased during exposure to low concentrations of
produced water There also was a significant up-regulation of the mRNA expression
of β-2-microglobulin, immunoglobulin-M light chain, and interleukins-1β and 8,
and down-regulation of interferon stimulated gene 15 at slightly higher exposure
concentrations. However, because toxic effects are directly linked to dosage and
exposure time, the ecological significance of laboratory biomarker studies is
questionable. Factors such as fish movement and contaminant uptake/elimination are
not taken into account and alkylphenol concentrations in seawater near platforms
usually are below the limits of detection.
34
Andrews et al. (this volume) conducted chronic toxicity studies on adult cunner
(Tautogolabrus adspersus) and juvenile cod (Gadus morhua) exposed to produced
water from the Hibernia field on the Grand Banks. The health effect indicators
studied included fish and organ condition, visible skin and organ lesions, levels of
MFO enzymes, haematology (differential white blood cell counts), a variety of
histological indices in liver and gills and vitellogenin - a biomarker for “estrogenic”
endocrine disruption. The fish were dosed every 2-3 days for 6-8 h with 1-2 parts-
per-thousand of produced water. Exposures were carried out for a 3-month period.
No changes in various indices were noted with the exception of red and white pulp
in the spleen which is associated with red and white blood cells respectively.
Lymphocyte levels also were depressed in the blood (Payne et al., 2005). Similar
results were obtained with juvenile cod, except that haematological effects were not
observed in cod. However, elevated levels of MFO and vitellogenin were recorded
in the exposed fish (Payne et al., 2005).
Gagnon (this volume) and Codi King et al. (this volume) performed similar studies
with tropical fish in the vicinity of offshore production platforms on the Northwest
Shelf of Australia. Gagnon (this volume) collected three species of fish from surface
waters near three production platforms or FPSOs discharging large volumes of
produced water and measured several physiological parameters and biomarkers in
the fish. Condition factor was slightly reduced in fish from one platform and liver
somatic index was elevated in fish captured at two of the platforms. EROD activity
and incidence of DNA damage were high at one facility discharging high volumes of
produced water. Stress proteins HSP70 were elevated in fish collected at all three
facilities. High concentrations of naphthalene and pyrene metabolites were detected
in the bile of fish collected at all three facilities. Gagnon (this volume) concluded
that the chemical composition and discharge rate of produced water affected the
biological responses observed in resident fish.
Codi King et al. (this volume) deployed juvenile Spanish flag snapper (Lutjanus
carponotatus) in cages for ten days at ~200 m, ~1000m, and a distant reference
location from a platform on the Northwest Shelf of Australia. The produced water
contained 10-14 ppm monocyclic aromatic hydrocarbons, about 2.6 ppm phenol and
C1- and C2-phenols, and about 1 ppm total PAH, mostly naphthalenes and
phenanthrenes. None of these chemicals were detected in bulk water samples
collected from the deployment sites; however, low concentrations of several PAH
were detected in SPMDs deployed for ten days with the fish cages. A large number
of biomarkers were evaluated for evidence of response to chemicals in the produced
water. Bile metabolites, CYP1A, CYP2K- and CYP2M-like protein, and liver
histopathology provided evidence of exposure and effects after 10 days at the two
sites near the platform, in comparison to results for fish at the reference site.
Hepatosomatic index, cholinesterase, and total cytochrome P450 were not
significantly different in fish from the three sites, whereas EROD activity was
inconclusive. Principal component analysis (PCA) validated that the most useful
diagnostic tools for assessing exposure to, and effects of exposure to produced water
in snapper were the CYP proteins.
35
Payne et al. (this volume) studied the toxicity to cunner of particulate barite (BaSO4)
that forms when produced water containing a high concentration of dissolved barium
mixes with seawater, rich in inorganic sulfate (Trefry and Trocine, this volume).
Cunner were exposed on a weekly basis for 40 weeks to 200 g “clouds” of
microparticulate barite in a 1,800-L tank (nominal concentration, 111 mg/L BaSO4,
compared to a seawater solubility of 0.08 mg/L). Barite that accumulated on the
bottom of the tank was not removed. Fish survival and indices of fish health, as
assessed by fish and organ condition as well as detailed histological studies on liver,
gill and kidney tissue did not differ between control and experimental groups.
However, slightly elevated activity of EROD was observed in the exposed fish. The
cause of the induction is not known.
A preliminary study was performed with scallops. The scallops were exposed to
2000 ppm produced water from the Hibernia field every 2-days for a period of ~ 4
months. No differences in mortality or condition indices were observed between the
control and exposed groups. A similar long-term study with mussels found no effect
of produced water on mortality (J. Payne, DFO – personnel communication).
6 Ecological Risk of Produced Water Discharges
The toxicity and ecological effects of a complex chemical mixture, such as produced
water, to marine organisms and communities is a product of its chemical
composition, environmental fates of each component in the mixture, and the relative
toxicities of each component and its degradation products. Increasingly complex
and sophisticated fate and effects models, such as DREAM, are being developed to
predict the long-term effects and ecological risks of produced water discharges to
different marine environments (Reed and Rye, this volume). Risk assessments can
be performed with the DREAM model by two approaches: 1) a dose-effect risk
assessment model, in which the dose is measured as the concentration of each
component in the ambient water or tissues of the target marine animals, and 2) the
determination of an environmental impact factor (EIF). The assessment is based on
the ratio of the predicted environmental concentration (PEC) to a predicted no-effect
concentration (PNEC), known as the PEC/PNEC ratio (Karman and Reerink, 1998).
This can be followed up with the calculation of an environmental impact factor
(EIF) to be used for produced water impact reduction, management and regulation
(Johnsen and Frost, this volume). The EIF is a measure of the volume of seawater
that contains high enough concentrations of produced water chemicals to exceed a
pre-determined risk criterion. The EIF provides a regional-scale, quantitative
estimate of the potential ecological risks to marine organisms of produced water
discharges (Johnsen et al., 2000).
The Norwegian oil and gas industry advocates ecological risk assessment as the
basis for managing produced water discharges to the North Sea. Neff et al. (2006)
36
compared estimates of ecological risks of PAH from produced water to water-
column communities based on data on hydrocarbon residues in soft tissues of blue
mussels deployed for a month near offshore platforms and based on predictions of
the DREAM model. The study was performed near produced water discharges to the
Tampen and Ekofisk regions of the Norwegian Sector of the North Sea. Because
PAH are considered the most important contributors to the ecological hazard posed
by produced water discharges, comparisons focused on this group of compounds.
In the DREAM model, predicted environmental concentrations (PECs) for three
PAH fractions were estimated in the three-dimensional area around the produced
water discharge. Predicted no effects concentrations (PNECs) for each fraction were
based on the chronic toxicity of a representative PAH from each fraction divided by
an application factor to account for uncertainty in the chronic toxicity value. The risk
characterization ratio (RCR) is the sum of the ratios of PEC estimated by the
DREAM to the PNEC for several PAH groups (Figure 9). The hazard index (HI) is
the sum of the ratios of the measured concentrations of individual PAH in the water
to an equilibrium partitioning/toxicity-based PNEC value for each target PAH. A
risk value of 1 or higher indicates a possible risk to the health of marine organisms
from the site.
The deployed mussel approach is based on PECs of individual PAH, estimated from
PAH residues in mussels that had been deployed at different distances from
produced water discharges, and PNECs based on a Kow regression model. The
mussel method gave much lower estimates of ecological risk than the DREAM
method (Figure 9). The differences are caused by the much lower PNECs used in
DREAM than derived from the regression model, and by the lower concentrations of
aqueous PAH predicted by DREAM than estimated from PAH residues in mussel
tissues. However, the two methods rank stations at different distances from
produced water discharges in the same order and both identify two- and three-ring
PAH as the main contributors to the ecological risk of PAH in produced water
discharges. Neither method identifies a significant ecological risk of PAH in the
upper water column of the oil fields. The DREAM model may produce an overly
conservative estimate of ecological risk of produced water discharges to the North
Sea, because of the extremely conservative PNEC values for PAH fractions.
Myhre et al. (2004) have studied the reproductive effects of alkylphenols (APs) on
fish stocks in the North Sea using the DREAM model. The fish stock distributions
(cod, saithe and haddock, from the international bottom trawl surveys (IBTS)
database) and a PNEC for APs of 4 ng/L were used as the basis data for the
calculations of effects from the combined produced water discharges from three
major Norwegian oil fields (Tampen, Ekofisk and Sleipner). The total amount of
APs>C4 discharged from all the oil installations was estimated to be 25.6 kg per day,
dissolved in 364,300 m3/day of produced water (~7 µg/L total >C4-APs). DREAM
predicted that none of the fish accumulated APs to concentrations above the critical
body burden of 2 μg/kg in any of the simulations. The highest accumulated body
burden in any of the fish was 0.09 μg/kg. Myhre et al. (2004) concluded that the
overall results of the simulations with DREAM show that there is not a significant
37
ecological risk from >C4-alkylphenols in produced water discharges to the
Norwegian North Sea.
The accuracy of contaminant risk assessment models is dependent on the
identification and quantification of the various chemicals that induce the toxic
effects. Unfortunately, the causative agents of toxicity in the most toxic produced
waters are not known. Toxic responses may be linked to the extremely high total
dissolved solids (salinity) concentrations, altered ratios of major seawater ions, and
elevated concentrations of ammonia in some Gulf of Mexico produced waters
(Moffitt et al., 1992). Salinity and ion ratios quickly return to those in seawater
following ocean discharge of produced water and ammonia evaporates or degrades
rapidly. Thus, these contaminants of concern within the produced water discharge
stream rarely cause acute toxicity responses in the field.
Bacteria have very short generation times and respond rapidly to environmental
changes. Because bacteria are involved in primary production processes including
the production of organic carbon, nutrient cycling and the biodegradation and
biotransformation of contaminants, their use has been recommended for
environmental effects monitoring programs (Lee and Tay, 1998; Wells et al., 1998).
Studies (Anderson et al., 2000; Lee et al., this volume) with naturally occurring
bacteria have indicated the potential for produced water to both inhibit (short term
exposure at high concentrations) and enhance bacterial growth (lower concentrations
over an extended period).
Chemical reactions that occur following the release of hypoxic produced water into
well-oxygenated open ocean water alter the toxicity of produced water over time
following its discharge (Lee et al., 2005a; Binet et al., this volume). The significance
of this process is clearly illustrated in controlled dose-response experiments using
natural microbial populations as the test organisms (Figure 10). A typical toxicity
dose-response curve (initial increase in productivity at low concentrations of
produced water due to addition of nutrients, followed by inhibition above a threshold
value) is observed with fresh produced water. However, following aeration for 44
hours (to simulate equilibration in the ocean following discharge) additions of the
produced water over the same concentration gradient elicited a stimulatory response.
The difference is attributed to the loss of low molecular weight hydrocarbons,
precipitation of redox- or sulfate-sensitive metals, such as barium, iron, and
manganese that sequester toxic metals, and photo oxidation of some organic
produced water chemicals. The results summarized in Figure 10 imply that accurate
comparisons of toxicological studies with similar end-points (e.g., LC50) cannot be
made unless sample collection, handling, and storage protocols are standardized
prior to toxicity testing.
In a modeling study to assess potential perturbations in food web structure and
energy flow due to the discharge of produced water, Rivkin et al. (2000) predicted
significant increases in productivity and sedimentation fluxes over large spatial
domains in response to ammonia and dissolved organic carbon from produced water.
However, at current discharge rates, the effects of produced water discharges may be
38
limited. Yeung et al. (this volume) monitored changes in indigenous microbial
community structure in response to produced water discharges from an offshore
platform on the Grand Banks of Canada by denaturing gradient gel electrophoresis
(DGGE). The DGGE results showed that the produced water did not have a
detectable effect on microbial community structure in the surrounding water.
Cluster analysis showed a >90% similarity for all near surface water (2 m) samples,
~86% similarity for all the 50 m and near bottom (NB) samples, and ~78% similarity
for the whole water column from top to bottom across a 50 km range, based on two
consecutive yearly sampling events. However, there were clear differences in the
composition of the bacterial communities within the produced water compared to
seawater near the production platform (~50% similar), indicating that the effect of
produced water may be restricted to the region immediately adjacent to the platform.
Members of the genus Thermoanaerobacter and of the Archaea genera
Thermococcus and Archaeoglobus were identified as significant components of the
produced water. These particular signature microorganisms could become useful
markers to monitor the dispersion of produced water into the surrounding ocean.
7 Environmental Effects Monitoring and Research Needs
This overview has highlighted our advances in scientific knowledge, during the past
decade regarding the composition, environmental fate and biological effects of
produced water discharges to the ocean.
The principal alternative to ocean discharge of treated produced water is
underground injection. However, the feasibility of this practice at offshore
installations is dependent on a number of site-specific factors including access to a
suitable disposal geologic formation, chemical interactions that may result in
precipitates that may plug the receiving formation, and cost. Furthermore, the net
environmental benefit of reinjection must be considered. On the basis of energy
requirements, it is estimated that 2.6–4.3 g of CO2 emissions are produced for each
liter of produced water reinjected into a sub-surface well (Shaw et al., 1999).
The general consensus of the 2007 International Produced Water Conference (St.
John’s, Newfoundland, Canada) was that any effects of produced water on
individual development sites in the open ocean are likely to be minor. The toxicity
threshold limits for acute effects are not likely to occur beyond the immediate
vicinity of the discharge pipe due to the effectiveness of natural dispersion processes
driven by tides and currents. However, unresolved questions regarding aspects of
produced water composition and its fate and potential effects on the ecosystem
remain. The chronic effects on important marine communities may become evident
only after monitoring several life stages, generations of keystone species, or long-
term ecological effects. It is important to acknowledge the consequences of long-
term effects from offshore oil and gas facilities that may have a 15-20 year life-
cycle. Furthermore, cumulative effects linked to future expansion of production
39
operations must be considered. It is evident that additional information is needed to
improve the accuracy of existing risk assessment models for produced water
discharge. Multidisciplinary scientific studies are needed under an ecosystem based
management (EBM) approach to provide information on the environmental fates
(dispersion, precipitation, biological and abiotic transformation) and effects of
chronic, low-level exposures of the different chemicals in produced water.
Numerical models need to be improved to better predict the fate and effects of
chemical constituents in produced water plumes that are rapidly dispersed. There is a
need to develop improved sample recovery and analytical techniques to support
model validation needs. At present, many of the potential contaminants of concern in
produced water cannot be detected in the open ocean environment with standard
analytical protocols. The future development of high efficiency, cost-effective
produced water treatment technologies is dependent on the identification and
monitoring of the primary target constituents of environmental concern (e.g. PAH,
phenols, naphthenic acids, metals) in produced water and the produced water plume.
Interpretation of ecological risk from biological effects studies based on biomarker
techniques remains a challenge. Biomarkers may be used to indicate that: 1) an
organism has been exposed to a specific chemical or group of chemicals, 2) an
organism is affected by a contaminant and is responding to it, 3) the organism has
been injured. However, as discussed by Gray (2002) in an editorial comment
entitled, “Perceived and real risks: Produced Water from oil extraction,” the question
is, “What is the risk to populations in the field?”
For a comprehensive protection plan, there is a need to support the development of
improved monitoring protocols to provide early warning of any potential problems
related to sediment and water quality (such as primary productivity), fish quality,
and fish health. Development of real-time monitoring systems (contaminant-specific
sensors and data transfer technologies) will enhance our capacity to manage the
ocean and its living resources. In consideration of natural perturbations currently
occurring within the ocean (climate change, for instance) and the impacts potentially
associated with other users of the oceans (marine transport, fisheries, etc.), an
ecosystem based integrated management approach must be taken to fully evaluate
the risks of produced water discharge into the oceans.
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Figure Captions
Fig. 1 Concentrations of individual PAH in produced water from two production
facilities in the U.S. Gulf of Mexico (based on data in Neff et al., this volume)
50
51
Fig. 2 Concentrations of phenol and C1- through C9-alkylphenol congener groups
in produced water discharged from three production facilities on the Norwegian
outer continental shelf (data from Boitsov et al., 2007)
Fig. 3 Mean concentrations of total petroleum hydrocarbons (TPH) in produced
water discharged offshore in several regions of the world, based on reports to OGP
(2004) – no data were available for the former Soviet Union
Fig. 4 Diagram of the processes governing the physical, chemical, and biological
behavior of target chemicals from produced water in DREAM (Reed and Rye, this
volume)
Fig. 5 Concentrations (ng/L = parts per trillion) of three PAH fractions in surface
waters at five distances down-current from production facilities in the Ekofisk
field in the Norwegian North Sea, based on measured concentrations in tissues of
deployed mussels and predictions of the DREAM model; concentrations in water
at R2 are considered background for the southern North Sea (data from Durell et
al., 2006, and Neff et al., 2006)
Fig. 6 Time series of modeled (DREAM) total PAH (naphthalenes through 5-ring
PAH) concentrations at Station S4 and Station S6, 0.5 and 5 km downcurrent,
respectively from a produced water discharge in the Ekofisk Field (from Durell et
al., 2006)
Fig. 7 Hepatic EROD activity in juvenile cod after five to six weeks deployment
in cages at different distances from production platforms in the Statfjord oil field
in the North Sea (from Førlin and Hylland, 2006)
Fig. 8 Alkylnaphthalene metabolite concentrations in the bile of cod deployed for
five to six weeks in cages at different distances from production platforms in
Statfjord oil field (from Førlin and Hylland, 2006)
Fig. 9 Risk characterization ratios (RCRs) and hazard indices (HIs) for total
polycyclic aromatic hydrocarbons (PAHs) at different distances from produced
water discharges in the Ekofisk field in the North Sea; stations are identified in the
x-axis by their distance (km) and direction (U = up-current; D = down-current; P =
parallel to) relative to the nearest produced water discharge (Neff et al., 2006)
Fig. 10 Toxicity dose-response data illustrating the effect of produced water
recovered from the Terra Nova FPSO, the Grand Banks, on microbial productivity
in natural seawater (measured by 3H-thymidine uptake into DNA); identical
samples were evaluated under identical conditions, immediately after collection
(Fresh) and after aeration (Aerated) for 44 hours (Lee et al., 2005)
... A pseudo-steady state Ecorr had not been attained 512 after 28 days. PW often contains various organic compounds, including hydrocarbons and organic 513 acids [37]. Additionally, PW typically contains inorganic compounds such as salts, minerals, and metal 514 ...
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