Assessment of the 

Environmental Effects 

Associated With Wooden 

Bridges Preserved With 

Creosote, Pentachlorophenol, 

or Chromated Copper 

Arsenate 

Kenneth M. Brooks 

United States

Department of 

Agriculture 

Forest Service 

Forest 

Products 

Laboratory 

Research 

Paper 

FPL−RP−587

In cooperation 

with the 

United States 

Department of 

Transportation 

Federal 

Highway 

Administration 
Abstract 

Timber bridges provide an economical alternative to concrete 

and steel structures, particularly in rural areas with light to 

moderate vehicle traffic. Wooden components of these 

bridges are treated with chromated copper arsenate type C 

(CCA), pentachlorophenol, or creosote to prolong the life of 

the structure from a few years to many decades. This results 

in reduced transportation infrastructure costs and increased 

public safety. However, the preservative used to treat the 

wooden components in timber bridges is lost to the environment in small amounts over time. This report describes the 

concentration of wood preservatives lost to adjacent environments and the biological response to these preservatives 

as environmental contaminants. Six bridges from various 

states were examined for risk assessment: two creosotetreated bridges, two pentachlorophenol-treated bridges, and 

two CCA-treated bridges. In all cases, the largest bridges 

located in biologically active environments associated with 

slow-flowing water were selected to represent worst-case 

analyses. Sediment and water column concentrations of 

preservative were analyzed upstream from, under, and downstream from each bridge. The observed levels of contaminant 

were compared with available regulatory standards or 

benchmarks and with the quantitative description of the 

aquatic invertebrate community sampled from vegetation and 

sediments. Pentachlorophenol- and creosote-derived polycyclic aromatic hydrocarbons (PAHs) were not observed in 

the water near any of the selected bridges. 

September 2000 

Brooks, Kenneth M. 2000. Assessment of the environmental effects associated with wooden bridges preserved with creosote, pentachlorophenol, or 

chromated copper arsenate. Res. Pap. FPL–RP–587. Madison, WI: U.S. 

Department of Agriculture, Forest Service, Forest Products Laboratory. 

100 p. 

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public from the Forest Products Laboratory, One Gifford Pinchot Drive, 

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endorsement by the U.S. Department of Agriculture of any product or 

service. 

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However, low levels of PAHs were observed in the sediments under and immediately downstream from these 

bridges. Pentachlorophenol concentrations did not approach 

toxicological benchmarks. Sediment concentrations of naphthalene, acenaphthylene, and phenanthrene exceeded the 

probable effect level. Metal levels at the bridges treated with 

CCA were less than predicted effect levels, in spite of questionable construction practices. Adverse biological effects 

were not observed in the aquatic invertebrate community or 

laboratory bioassays conducted on water and sediments 

sampled at each of the bridges. Results of this study reveal 

the need to follow the construction information found in Best 

Management Practices for the Use of Treated Wood In 

Aquatic Environments published by Western Wood Preservers Institute. Regulatory benchmarks used in risk assessments 

of this type need to be indexed to local environmental conditions. The robust invertebrate communities associated with 

slow-moving streams over soft bottoms were not susceptible 

to the concentrations of PAHs that would be expected to 

affect more sensitive taxa, which typically are located in 

faster moving water over hard bottoms. Contaminants released from timber bridges into these faster systems (where 

more sensitive taxa are located) are significantly diluted and 

not found at biologically significant levels. 

Keywords: timber bridges, preservative, CCA, creosote, 

pentachlorophenol 

Acknowledgment 

We acknowledge the support of the U.S. Department of 

Transportation, Federal Highway Administration, for this 

research project. The author expresses sincere appreciation 

to Jean Livingston of the Forest Products Laboratory 

for her detailed editing of this report and Arthur Frost for his 

meticulous identifications of freshwater invertebrates. 
Contents 

 Page

Introduction............................................................................1

Background............................................................................3

Previous Treated-Wood Assessments....................................4

Creosote Evaluation...........................................................4

Creosote Mesocosm Studies ..............................................5

Metal Uptake .....................................................................5

Wetland Boardwalks..........................................................6

Materials and Methods...........................................................7

Site Selection .....................................................................7

Sample Location ............................................................... 8 

Data Analyses ....................................................................8

Physicochemical Characterization of Water Column.........9

Physicochemical Characterization of Sediments..............10

Characterization of Biological Response .........................10

Cleaning of Sample Containers and Equipment...............10

Sample Collection and Field Processing..........................10

Sample Documentation and Handling..............................10

Sediment and Water Analyses for Preservatives............. 11 

Sediment and Water Analyses for Conventional 

Parameters........................................................................11

Quality Assurance Requirements and Data Qualifying 

Criteria .............................................................................12

Sediment Grain Size and Total Volatile Solids............12

Infaunal Analysis .........................................................12

Results and Discussion.........................................................12

Creosote-Treated Bridges................................................12

Sources of PAH ...........................................................13

Observed Levels of PAH .............................................14

Fate of PAH .................................................................14

Bioconcentration, Bioaccumulation, and 

Biomagnification..........................................................15

Creosote and PAH Toxicity.........................................17

Potential for Human Pathology Associated 

With PAHs...................................................................22

 Page 

Sediment Quality Benchmarks For PAHs................... 22

Site Selection ...............................................................23

Bridge 146 ...................................................................23

Bridge 148 ...................................................................31

Risk Assessment Summary of Bridges 146 and 148 ....36

Penta-Treated Bridges......................................................37

Sources.........................................................................37

Environmental Chemistry ............................................38

Fate ..............................................................................38

Metabolism ..................................................................44

Bioconcentration and Bioaccumulation .......................45

Toxicity........................................................................47

Loss of Penta................................................................52

Dioxins in Penta...........................................................53

Results From Penta-Treated Cougar Smith Bridge ......53

Results From Upper Dairy Creek Bridge.....................57

CCA-Treated Bridges......................................................63

Background Levels and Sources..................................63

Cycling and Fate ..........................................................65

Bioconcentration, Bioaccumulation, and 

Biomagnification..........................................................67

Toxicity to Aquatic Fauna and Flora ...........................68

Regulatory Benchmarks...............................................70

Summary of Sources and Toxicity in Aquatic 

Environments...............................................................71

Results From Horseshoe Bayou Bridge .......................71

Results From Fountains CCA-Treated Bridge .............80

Conclusions..........................................................................85

References............................................................................86

Appendix A—Test America Reporting Limits for 

Polycyclic Aromatic Hydrocarbons (PAHs)........................96

Appendix B—Freshwater Taxonomy Codes........................97

Appendix C—Marine Taxonomy Codes............................100

Assessment of the Environmental Effects 

Associated With Wooden Bridges Preserved 

With Creosote, Pentachlorophenol, or 

Chromated Copper Arsenate 

Kenneth M. Brooks, Owner and Principal Scientist

Aquatic Environmental Sciences, Port Townsend, Washington

Introduction 

Wood is a renewable resource that can be modified to produce bridge structures that are strong and relatively light 

weight, compared with steel or concrete, and can be aesthetically pleasing. Wood preservatives are used to increase the 

life span of wooden bridge structures from a few years to 

many decades. Wooden bridges, preserved with a variety 

of chemicals, have been used widely throughout the 

United States and Canada. They are particularly valuable for 

small- to moderate-sized pedestrian and vehicular bridges in 

rural areas. 

Timber bridges are typically treated with pentachlorophenol 

(penta) or creosote. These oilborne preservatives are preferred for long spans where the wood must remain flexible. 

However, wood treated with waterborne preservatives, 

represented primarily by ammoniacal copper zinc arsenate 

(ACZA) and chromated copper arsenate type C (CCA), is 

also being used in the construction of wooden bridges. A 

preliminary survey conducted in support of this study indicated that creosote is used primarily in the Northeast and 

Midwest, penta and creosote in the Pacific Northwest, and 

CCA throughout the South. These are qualitative trends, and 

examples of the use of each type of preservative can be found 

throughout the United States. 

This study began with an extensive literature search. That 

search did not uncover any reports documenting a loss of 

biological integrity associated with timber bridges constructed of treated wood. However, a perceived environmental degradation is often associated with these structures. 

Figure 1, a photograph taken in 1995, shows a bridge crossing Narragansett Bay. The bridge was constructed of laminated beams that began dripping creosote during a series of 

hot summer days. Figure 2 shows the accumulation of 

dripping creosote on rocks under the bridge. 

Krahn (1987) reported the leaching of copper and arsenic 

from ammoniacal-copper-arsenate- (ACA-) treated wood 

used as bridge decking and support timbers in British Columbia, Canada. Initial treatment of this wood did not meet 

Canadian Wood Preservers’ Association (CWPA) standards, 

and the wood was retreated. Krahn’s (1987) investigation 

suggested that the wood was immediately placed in service 

following the second treatment and that fixation of the metals 

had not occurred. Krahn (1987) discussed the potential for an 

avoidance reaction in salmonids at the copper levels observed in the stream but did not document actual environmental degradation. 

The literature is rich with descriptions of metal and hydrocarbon pollution associated with transportation systems. 

However, not one document was uncovered that describes 

the loss of wood preservatives from timber bridges. Therefore, this study breaks new ground in understanding the 

environmental risks associated with these structures. 

In 1997, the USDA Forest Service contracted with Oregon 

State University and Aquatic Environmental Sciences to 

conduct a three-phase study (contract RJVA–2828). The 

purposes of this study were to 

• assess the environmental response associated with existing 

timber bridges, 

• determine the loss of various types of preservatives from 

overhead bridge structures, and 

• develop a computer model to assist the Forest Service in 

understanding the site-specific environmental risks associated with proposed timber bridge construction. 
2 

Figure 2—Accumulated creosote from weeping pressure-treated timbers used in bridge construction.

Figure 1—Creosote loss from treated glulam beams used in constructing a highway bridge across 

Narragansett Bay. 
3 

Background 

Transportation systems are inherently dirty. Polycyclic 

aromatic hydrocarbons (PAHs) are released from many 

sources, including asphalt, lubricating oils, and gasoline or 

diesel engine exhaust. Asphalt contains significant concentration levels of PAH, including the known carcinogen, 

benzo(a)pyrene at 6 μg/g (Machado and others 1993). In 

Germany, Muench (1992) observed roadside levels of PAHs 

at 6.6 to 9.8 μg/g (dry soil weight) and zinc levels of 58 to 

330 μg/g. These levels approach those at which adverse 

biological effects could be anticipated in particularly 

sensitive aquatic environments. 

Elevated concentrations of heavy metals, including zinc and 

copper, were found within 100 m of roads with traffic volumes exceeding 1,000 vehicles per day (Hoedrejaerv and 

others 1997). Mean storm water levels of zinc and copper 

can exceed surface water quality discharge standards 

(Sasalone and Buchberger 1997). In Moscow, Lepneva and 

Obukhov (1990) observed copper at 80 μg/g and zinc at 

230 μg/g in roadside soils. These levels were four to six 

times greater than background levels. 

Sediment levels of contaminants in aquatic environments 

adjacent to roadways can be significantly elevated. Faganeli 

and others (1997) documented the uptake of road surface 

contaminants, including PAH, zinc, and copper, by plants, 

bivalves, and gastropods in the Gulf of Trieste. However, the 

levels of PAHs and metals were not sufficiently high to 

adversely effect biota (Faganeli and others 1997). The 

bioavailability of PAHs and their ability to influence biological systems are dependent on a number of environmental 

physicochemical parameters, particularly the concentration of 

organic carbon that binds PAH (Johnsen 1987, Brooks 

1997b). For example, Paine and others (1996) examined 

sediment PAH levels and biological responses associated 

with an aluminum smelter. Levels of 10,000 mg PAH per 

kilogram of dry sediment were observed in some samples 

within 1 km of the smelter. However, most samples revealed 

sediment PAH concentrations of about 150 mg/kg. Despite 

repeated sampling for several years, the authors found no 

sediment toxicity associated with four laboratory bioassay 

test animals and minimal evidence of toxicity in the resident 

benthic invertebrate community. The authors concluded that 

the lack of observed toxicity was associated with the reduced 

bioavailability of PAHs. 

Storch and others (1990) found that Chautaugua Lake Bridge 

runoff was toxic to young-of-the-year sunfish (Lepomis 

macrochirus). The authors suggested that sodium chloride 

(used to de-ice roads) was a major contributor to the toxicity 

but that concentration levels of zinc and cadmium were also 

found at toxic levels. Baekken (1994) compared metal and 

PAH levels in a small reference lake with those in a similar 

lake adjacent to a highway. He found that cadmium and zinc 

concentrations in bivalves were two to three times greater 

next to the highway when compared with the reference lake. 

The concentration levels of other metals (copper, mercury, 

nickel, lead) and PAHs in bivalves were low in both lakes, 

presumably close to background levels. The diversity and 

abundance of benthic communities were reduced on the 

highway side of the polluted lake, suggesting effects associated with highway pollutants. Yousef and others (1985) 

observed elevated levels of metals, including lead, copper, 

chromium, iron, nickel, cadmium, and zinc, in lake water and 

sediments associated with a highway bridge. The authors 

reported that 95% to 98% of the metals associated with the 

road and bridge were found in the sediments. Marsalek and 

others (1997) observed high levels of metals and PAHs in 

runoff from the Skyway Bridge in Burlington, Ontario. The 

observed levels in water and sediments are summarized in 

Table 1. The authors noted that sediment levels of metals 

were considered “grossly polluted” according to the Ontario 

Ministry of environment and energy guidelines for sediment 

quality. 

A large proportion of the contaminants entering aquatic 

environments appears to be sedimented at short distances 

from the point of introduction. Schiffer (1989) found that 

median zinc levels decreased from 75 μg/L at a storm water 

inlet to 20 μg/L at a distance of 30.5 m from the inlet. Therefore, the magnitude of the disturbance associated with transportation systems is partially dependent on the size (particularly the width with respect to the roadway) of the water 

body. Obviously small, stagnant ponds and narrow, slowmoving streams flowing parallel to major transportation 

routes would likely be more stressed than would a large, 

swift river flowing under a highway bridge. 

O’Malley and others (1996) used molecular abundance and 

carbon isotope measurements to partition four- and five-ring 

PAHs by source in Saint John’s Harbor, Newfoundland, 

Canada. Their mixing model suggested that approximately 

50% to 80% of the PAH input to the harbor was of combustion origin, likely dominated by vehicular emissions carried 

by surface runoff from the city of Saint John’s. In addition, 

direct petroleum-related contributions, dominated by crankcase oil, accounted for the remaining 20% to 50% of the total 

PAH input. 

Gravel roads can result in the deposition of significant 

amounts of sharply divided sediments in rivers and streams, 

Table 1—Metal and PAH levels observed in storm water 

run-off from the Skyway Bridge in Burlington, Ontarioa 

Contaminant Water (μg/L) Sediments (μg/g) 

Zinc 337 997 

Copper 136 314 

14 PAH 0.015 to 0.500 

a

Marsalek and others 1997. 
4 

with adverse effects on fish and invertebrate habitat (Cedarholm and Salo 1979, Cedarholm and others 1982). Road 

traffic tends to break rock into sharp fragments that can cause 

injury to the gills of aquatic organisms, whereas natural 

sediments tend to be more rounded by abrasion, thus having 

a lower propensity to cause physical injury. 

The point in this discussion is that transportation systems, 

including the roads that are carried over water bodies by 

bridges, are the source of many contaminants, including the 

metals used to preserve wood (copper, zinc, and chromium) 

and the PAHs contained in creosote. Unless an attempt is 

made to control these sources that are not associated with 

treated wood, they could significantly confound the results of 

a study intended to evaluate the environmental effects associated with preservative lost from treated wood. 

Previous Treated-Wood 

Assessments 

An extensive literature search did not reveal any reports that 

discussed environmental effects associated with timber 

bridges. Brooks (1995a,b, 1996, 1997a–c, 1998a,b) published risk assessment guides and computer models for assessing environmental risks associated with wood treated 

with creosote, CCA, Ammoniacal Copper Quat, type B 

(ACQ), and ACZA and used in aquatic environments. These 

documents focus on the loss of preservative from immersed 

wood and do not adequately address the loss of preservative 

constituents from overhead structures. 

The magnitudes and characteristics of losses of wood preservatives from overhead bridge structures into underlying 

aquatic environments are a function of numerous variables, 

including wood type, preservative, retention rate (how much 

preservative is in the wood), and production methods. These 

variables can be controlled by the engineer, specifier, and 

wood treater. In addition, loss of preservative is a function of 

rainfall quantity and pH for waterborne preservatives and 

wood temperature for oilborne preservatives. 

Goyette and Brooks (1999) hypothesized that much of the 

sedimented PAH associated with creosote-treated piling 

results from solar heating of the wood above the water line. 

On very hot days, creosote-treated wood acts as a black 

body, and the temperature of the wood may be much greater 

than ambient air temperatures. Under these conditions, the 

creosote may migrate to the surface of the wood where blisters can form and pop, ejecting small particles of creosote 

away from the piling. Alternatively, the creosote may run 

down the piling or timber until it forms droplets that fall to 

the water. Drs. John Simonsen and Jeff Morrell, Oregon 

State University, are evaluating the magnitude and nature of 

preservative losses from overhead structures. When these 

data are available, it is anticipated that a risk assessment 

computer program will be developed and available in the 

year 2000. 

Several studies examining the environmental response to 

treated-wood structures have recently been reported. These 

studies are reviewed in the following sections. 

Creosote Evaluation 

Goyette and Brooks (1999) examined biological and physicochemical responses to the presence of two six-piling 

creosote-treated dolphins, with 2.4- to 3-m-wide bases, 

placed in a pristine marine environment. The study compared 

aged (8-year-old) and newly treated piling with an untreated 

Douglas-fir control and an open control (lacking any structure). Water depths were 8.7 m and currents were slow, 

with a maximum speed of 1.89 cm/s. Based on the first 

535 days of evaluation, the authors concluded the following: 

• Creosote contamination occurred primarily as minute tar 

droplets or particles within surface and subsurface layers 

of surficial sediments. Water concentration levels of total 

PAHs immediately adjacent to the down-current piling 

were in the 0.020 to 0.030 μg/L range, which is a level 

where no biological effects should be anticipated. 

• Sediment levels of PAH reached 18 μg/g dry sediment 

(range 5.0 to 30.0 μg/g) within half a meter of the dolphin’s perimeter and declined exponentially to levels of 

about 7.5 μg/g at a distance of 7.5 m. Sediment levels of 

PAH were close to background levels (0.2 μg/g) at distances beyond 7.5 m from the dolphin’s perimeter. 

• The Creorisk Model of Brooks (1997b) predicted about 

30% more sedimented PAH than was actually observed 

during the first 535 days of the study. 

• An extensive infaunal community analysis was conducted. 

This resulted in the identification of PAH-sensitive and 

PAH-tolerant taxa. However, significant adverse effects 

were not seen in any compartment of the infaunal community at any distance from the structures. 

• Laboratory bioassays included the amphipods Rhepoxynius 

abronius and Eohaustorius washingtonianus and solid and 

liquid phase tests using a marine bioluminescent bacterium 

(Vibrio fischeri), plus echinoderm sperm fertilization tests. 

The results of laboratory bioassays indicated low-level toxicity associated with Sooke Basin reference sediments. Reduced survival was observed in proximity to the untreated 

Douglas-fir piling, and increased toxicity was found close 

to the creosote-treated piling. Taken altogether, bioassay 

data indicated that sediments located within 0.65 m of the 

creosote-treated dolphin were toxic in laboratory studies. 

• Elevated levels of PAH were observed in blue mussels 

(Mytilus edulis) grown at various distances from the sixpiling dolphins during the first 14 days of the study. Tissue 

levels then declined to background concentration levels. 
5 

• Mussels raised in cages within 15 cm of the treated piling 

grew more slowly than did those grown further away. Survival was excellent in all cohorts (>98%), with no significant differences between stations. All mussel cohorts 

spawned normally, and no difference was observed in the 

survival or development of the larvae. 

• This site was selected because it was representative of a 

“worst case” project—slow currents, fine-grained sediments, and low sediment organic carbon, coupled with a 

healthy infaunal community. The authors concluded that 

even in these worst-case environments, creosote-treated 

piling would adversely affect biological resources only in 

the immediate vicinity (<0.65 m) of the structure. Therefore, they recommended that creosote-treated wood be a 

product that is managed, particularly in sensitive aquatic 

environments with very slow currents. 

• A comparison of individual PAH compound concentrations with Long and others (1995), Washington Sediment 

Quality Criteria, and the Canadian Interim Sediment Quality Guidelines suggested that phenanthrene posed the highest level of risk, followed by lower risks associated with 

fluorene and acenaphthene. Fluoranthene and chrysene 

posed the most significant risks associated with highmolecular-weight compounds. Only the proposed U.S. Environmental Protection Agency (EPA) sediment quality 

criteria for acenaphthene, phenanthrene, and fluoranthene 

were found to be underprotective. The Washington State 

Sediment Quality Criteria were most efficient in that they 

were protective of infauna but produced the fewest number 

(12) of false negatives (instances in which adverse effects 

were predicted but not actually observed). The threshold 

effects level (TEL) was least efficient, predicting adverse 

effects in 52 cases where not one was observed. The 

probable effects level (PEL) was next most efficient with 

21 false positives. 

Creosote Mesocosm Studies 

Bestari and others (1998) examined the environmental response of different numbers of newly treated creosote piling 

placed in 12,000-L freshwater microcosms. The authors 

concluded the following: 

• Water column concentration levels of PAH increased for 

7 days following installation of the piling, followed by an 

exponential decline to approximately background levels 

within 84 days. 

• The loss of PAHs from the water was not reflected in 

increased sediment concentration levels. 

• When these initial concentration levels declined, the authors concluded that “the accumulation of creosoteassociated PAHs from impregnated pilings in aquatic environments will likely be minimal and probably not pose a 

significant risk to aquatic biota” (Bestari and others 1998). 

Metal Uptake 

Weis and Weis (1992) examined the uptake of CCA metals 

from bulkheads on the Atlantic coast of the United States. 

They found that the green algae Ulva lactuca and Enteromorpha intestinalis contained elevated levels of metals when 

compared with the same species collected from reference 

areas. Significant mortality was observed in snails (Nassarius 

obsoletus) fed an exclusive diet of the algae from CCA 

treated docks. The snails and algae were housed in small 

static containers, which exacerbated metal exposure. Unfortunately, water column concentration levels of metal were not 

reported in this paper. Similarly, Weis and Weis (1993) 

examined copper levels in oysters residing on CCA-treated 

piling and bulkheads. They reported American oyster 

(Crassostrea virginica) wet tissue weight copper levels of 

12.59 μg/g from reference areas containing no CCA-treated 

wood, an average level of 27.05 μg/g in oysters growing on 

CCA-treated piling, and a level of 154.3 μg/g in oysters 

growing on CCA-treated bulkheads in residential canals. The 

National Academy of Sciences (NAS 1971) gives a bioconcentration factor of 5,000 for copper in marine mollusks. 

When coupled with the known metal loss rates from CCAtreated wood (Brooks 1996), elevated levels of copper 

should be expected in oysters growing on treated wood. To 

put the observed levels in proper perspective, note that Shuster and Pringle (1969) summarized historical reports of trace 

metal levels in the American oyster (Crassostrea virginica). 

They found an average of 144.8 μg Cu/g wet tissue weight, 

with a range of 6.83 to 600 μg/g. The mean oyster tissue 

copper level reported by Shuster and Pringle (1969) was 

5.35 times greater than the level observed in the same species 

of oysters growing on CCA-treated piling in the Weis and 

Weis (1993) study. The mean copper level reported by Weis 

and Weis (1993) for oysters growing on CCA-treated bulkheads (154.3 μg/g) was only 7% greater than the mean for all 

American oysters reported by Shuster and Pringle (1969). 

Similarly, Goyette (1975) found whole body copper levels of 

2,550 μg Cu/g wet oyster tissue in Crassostrea gigas collected from Howe Sound, British Columbia, near a mine site. 

Reference area tissue copper levels were 380 μg Cu/g wet 

tissue in this study. The point is that a loss of biological 

integrity cannot be inferred simply because copper (or any 

other possible pollutant) is significantly elevated. In this 

case, the elevated levels of copper observed in oysters growing on CCA-treated wood was well within the normal range 

of this metal observed throughout the East or West Coasts of 

the United States. 

Weis and Weis (1995) examined the benthic impacts of four 

CCA-treated wood structures in National Estuarine Research 

Reserves. Sediment quality benchmarks (Long and others 

1995, Suter and Tsao 1996, Jones and others 1997) are based 

on the fraction of a contaminant in the whole sediment 

weight, not on the percentage fines as reported by Weis and 

Weis (1995). Furthermore, the National Oceanic and 
6 

Atmospheric Administration (NOAA 1988) cautioned that 

evaluating the biological risks associated with metal concentration levels in only the fine fraction (<63 μm particle diameter) of sediments can lead to unacceptable bias in the 

interpretation, especially when fines represent <20% of the 

sediment matrix. In the Weis and Weis (1995) study, the 

percentage fines adjacent to bulkheads varied between 1.3% 

and 3.9% in sediments containing 5.0 to 25.0 μg/g copper in 

this fine fraction. Converting these levels to total sediment 

dry weight gives copper concentration levels of <1.0 μg/g at 

all distances from the bulkhead. These were exceptionally 

low values that did not approach NOAA’s effects range low 

of 34 μg/g dry sediment weight. Sedimented copper levels 

observed by Weis and Weis (1995) adjacent to CCA-treated 

bulkheads were lower by at least an order of magnitude than 

concentration levels at which any biological effect might be 

observed. Weis and Weis (1995) found that “Despite the fact 

that some (piling) were less than 1 year old, no clear accumulation of metals in sediments or organisms was noted; nor 

were any consistent effects seen on community parameters.” 

The authors concluded “pilings do not appear to produce 

significant impacts to the nearby environments, at least in 

well-flushed systems.” 

Adler–Ivanbrook and Breslin (1999) exposed blue mussels 

(Mytilus edulis) to wood treated to a preservative retention of 

40 kg/m3

 CCA for 9 months in 1994 and 3 months in 1995. 

They found few significant differences in condition index, 

dry tissue weight, or valve length between mussels grown 

adjacent to treated wood and control mussels. They found 

that the tissue levels of copper, chromium, and arsenic in 

both sea table experiments and field exposures were generally within normal ranges found in mussels from sites within 

Long Island Sound and around the United States. 

Wendt and others (1994) examined the concentration and 

biological effects of metals (copper, chromium, and arsenic) 

and organic compounds lost from treated-wood docks in 

South Carolina macrotidal creek systems. Copper, chromium, 

arsenic, and PAHs were measured in composite samples of 

surficial sediments and naturally occurring oyster populations 

(Crassostrea virginica) from creeks with high densities of 

docks and from nearby reference creeks with no docks. The 

results indicated that in macrotidal estuarine environments, 

wood preservative leachates from dock pilings had no 

acutely toxic effect on four common estuarine species nor did 

they affect the survival or growth of juvenile oysters for 

6 weeks. In some cases, metal leachates accumulated in 

sediments and oysters immediately adjacent to pilings but 

were not elevated elsewhere in the same creeks. 

Wetland Boardwalks 

Brooks (2000) examined the biological response to an extensive boardwalk system constructed in wetlands associated 

with the Wildwood Recreational Area on the western slope 

of Mount Hood in Oregon. Three spurs constructed of 

ACZA-, ACQ-, and CCA-treated woods were evaluated. 

Each spur was extended into aquatic areas of the wetland 

system. A mechanical control structure, constructed of untreated Douglas-fir, was constructed in a hydrologically 

isolated part of the wetland for comparison. Metal levels 

were measured in sediments and water and compared with 

characteristics of the resident aquatic invertebrate community 

before construction and at 15, 162, and 336 days following 

construction. Samples were collected at upstream control 

stations, at the edge of each structure, and at various distances downstream. 

The different structures contributed various levels of copper, 

chromium, arsenic, or zinc to both the water column and 

sediments. Based on leaching rates and toxic thresholds in 

water or sediments, copper lost from CCA-treated wood 

presented greater environmental risks than did either arsenic 

or chromium (chromium(III) or (VI)). The U.S. EPA 

(USEPA 1984) published water quality criteria for copper in 

freshwater: (1) short-term (acute) exposure and (2) long-term 

(chronic) exposure. These criteria, which are dependent on 

water hardness, are described in Figure 3. 

U.S. EPA acute copper criterion 

< 0.960 exp{0.9422[ln(hardness)]1.464} (1) 

Note: The acute copper criterion is a 1-h average not to be 

exceeded more than once every 3 years. 

U.S. EPA chronic copper criterion 

< 0.960 exp{0.8545[ln(hardness)]1.465} (2) 

Note: The chronic criterion is a 4-day average not to be 

exceeded more than once every 3 years. 

Freshwater sediment quality criteria have not been established for the metals contained in CCA. Until appropriate 

freshwater sediment quality standards are developed, the 

benchmarks given in Table 2 are proposed for evaluating the 

benthic effects associated with copper. The development of 

these benchmarks is described in detail by Brooks (2000). 

Maximum levels of dissolved or sedimented copper, chromium, or arsenic did not exceed a high no-effect concentration in the aquatic environment adjacent to the CCA-treated 

bridge and viewing platform in the Wildwood study. Biological endpoints evaluated in that study included species 

abundance, taxa richness, diversity (Shannon’s Index), and 

community evenness (Pielou’s Index). Adverse effects on 

aquatic invertebrates were not anticipated at the levels recorded in Wildwood, and adverse effects were not observed 

in invertebrate samples collected from vegetation, from 

replicated artificial substrate samples, or from sediments. 

Sediment concentrations adjacent to one of the other two 

treated-wood platforms reached concentration levels of 

200 μg Cu/g dry sediment. However, a reduction in the 
7 

Figure 3—U.S. EPA chronic and acute copper criteria for freshwater. The copper standard is 

presented in micrograms per liter and hardness values in milligrams of calcium carbonate per liter. 

Table 2—Recommended benchmarks for assessing 

environmental risks associated with sedimented 

copper lost from pressure-treated wood 

Sediment and water column characteristics 

Acceptable 

levels of 

sedimented 

copper 

(μg Cu/g 

dry 

sediment) 

Coarse-grained sediment (silt and clay <10%) 30 

Total organic carbon <0.2% 

Moderate to low pH (5.5 to 6.5) 

Low hardness and alkalinity (15 to 

25 ppm calcium carbonate) 

Intermediate sediments (silt and clay 

between 10% and 25%) 

55 

Total organic carbon between 0.2% 

and 1.0% 

Neutral pH (6.5 to 7.5) 

Moderate hardness and alkalinity 

(35 to 100 ppm calcium carbonate) 

Low-energy, well-buffered streams and lakes 

(fines >25%) 

100 

Total organic carbon >1.00% 

Greater than neutral pH (pH > 7.5) 

High hardness and alkalinity 

(>100 ppm calcium carbonate) 

biological endpoints measured in the study was not observed, 

even at these levels. This suggests that the benchmarks proposed in Table 2 may be overly conservative from an environmental point of view. 

Materials and Methods 

The following protocols were followed in conducting this 

study. 

Site Selection 

Primary producers of timber bridges were identified and 

interviewed to determine the types of preservatives commonly used. Based on that information, and in consultation 

with the Forest Products Laboratory (FPL), a decision was 

made to evaluate two creosote-treated bridges in the Midwest, two penta-treated bridges in the West, and two CCAtreated bridges in the southern tier of the United Sates. Producers of these bridges were requested to nominate bridges 

meeting the following site selection criteria: 

• CCA-treated bridge candidates should be constructed 

within 90 days of being evaluated to observe the early 

loss of metals. 

• Creosote- and penta-treated bridges should be 2 to 3 years 

old to allow time for preservative accumulation in 

sediments. 

0

10

20

30

40

50

60

70

Hardness (ppm CaCO3)

Copper criterion (μg/L)

Acute value 4.6 8.9 13.0 17.0 21.0 24.9 28.8 32.7 36.5 40.3 44.1 47.9 51.7 55.4 59.1 62.8

Chronic value 3.5 6.3 8.9 11.4 13.7 16.1 18.3 20.5 22.7 24.8 26.9 29.0 31.1 33.1 35.1 37.1

25 50 75 100 125 150 175 200 225 250 275 300 325 350 375 400
8 

• The stream should have healthy riparian (streamside) 

vegetation. 

• Bridges should be substantial in size and should carry 

vehicular traffic. A large pedestrian bridge would be acceptable but is not a first choice. 

• Bridges should be located in rural areas, away from other 

anthropogenic sources of metals, penta, or PAHs. 

• Bridges should have treated wood decks and, if possible, 

be supported on treated-wood piling that is immersed in 

water. 

• The stream should be relatively slow flowing, preferably 

<10 cm/s and with substantial flows >0.56 m3

/s (>20 ft3/s). 

• The stream should be perennial and contain freshwater, not 

seawater. 

• The stream bottom should be soft (sand, silt, and clay 

rather than coarse rock and cobble). This bottom structure 

would be expected where slow currents exist. 

• The stream should have a minimum depth of 30 cm. 

• Bridges should not be subject to abrasion by commercial 

boat traffic or barges. 

• The treated wood in bridges should be properly treated and 

should not be painted or wrapped 

Sample Location 

The adopted design relied on analysis of variance (ANOVA) 

and regression analyses to identify differences in sediment 

physicochemical and biological endpoints at upstream 

controls, directly under each bridge, and at various distances 

downstream. Sample station locations are described in 

Figure 4. Eighteen stations were evaluated at each bridge. 

Data Analyses 

The study design allowed for either a t-test or ANOVA 

testing that compared the replicated end points at reference 

stations located >30 m upstream with those collected under 

or immediately downstream from the drip line of the bridge. 

In addition, regression analysis was used to search for clines 

in biological and physicochemical endpoints from the upstream to downstream sample stations. The significance of 

the clines was determined by evaluating the statistical significance of coefficients on the various independent variables 

evaluated in the study. Correlation analysis, including principle components analysis, was used to evaluate the relationship between biological responses and physicochemical 

endpoints and between the bridge structure (distance) and 

the concentration of treated-wood preservatives observed in 

either the water or sediments. Fourth-order polynomial least 

squares fits to the biological endpoints were constructed. 

The shapes of these curves provide a qualitative assessment 

of the biological effects associated with each bridge. A guide 

to interpreting these graphs is provided in Figure 5.

 Figure 4—Generalized location of sample stations examined at each timber bridge evaluated in this study. 

River or stream current

Sample transect

+9.9 m +6 m +3 m +1.8 m +0.9 m +0.3 m 0 > - 9 m

 

Downstream Bridge Upstream (reference)

 Infaunal sample

 Bioassay sample

 Sediment and/or water physicochemical sample (sediment grain size, total volatile solids, 

preservative (Cu, Cr, As, PAH, or penta))































 
9 

Cause and effect relationships between environmental concentration levels of preservative constituents and biological 

endpoints were investigated using regression, correlation, 

and principle components analyses. In some analyses, a 

variable, such as copper concentration, would be considered 

independent (with respect to biological endpoints), and in 

another part of the analysis, the same variable would be 

dependent, such as the copper concentration with respect to 

distance from the bridge structure. 

The following biological endpoints were evaluated as dependent variables in this study: 

• Total species richness (total number of taxa) and Margalef’s (1958) Richness Index—Ludwig and Reynolds 

(1998) discussed the underlying assumption for a functional relationship, of the form S = k n , between the 

number of taxa S and the number n of organisms in the 

sample. This relationship was not found in this database, 

and valid conclusions were unlikely to follow from the 

analysis of Margalef’s Index. Ludwig and Reynolds (1998) 

noted that simple species richness S is a more appropriate 

endpoint when sample sizes are equal. This analysis evaluated the number of taxa S. 

• Total sample abundance is simply the number of organisms 

present in a single sample. This metric is useful in identifying environments where stress is severe enough to adversely affect all taxa, including those that are normally 

tolerant of the stressor. It is not a useful metric for assessing subtle effects that adversely impact intolerant species 

that are replaced, numerically, by more tolerant taxa. 

• Dominant and subdominant species abundance—Dominant 

taxa were identified as those that represented at least 1% of 

the total number of organisms in an individual bridge’s 

taxonomic database. Correlation analysis was used to examine the relationship between abundance of dominant and 

subdominant taxa and sediment or water column concentrations of copper, chromium, arsenic, zinc, penta, or 

PAHs as appropriate for the type of treated wood used in 

constructing the bridge. Highly dominant taxa that were 

not susceptible to the chemicals of concern were dropped 

from the database because their numerical strength would 

have masked effects on more sensitive, but less abundant, 

taxa. The resulting group of subdominant and more sensitive taxa was considered most likely to reveal adverse effects. 

• Shannon’s Index (Shannon and Weaver 1949) provides the 

average uncertainty per species in an infinite community of 

S taxa. The form of the index used in this analysis is given 

by Equation (3). The value of Shannon’s Index is zero 

when a single species is present. The value is maximized 

when there are a large number of equally represented species and is reduced in communities dominated by a few 

highly abundant species. The value of Shannon’s Index in 

a sample containing 20 taxa equally represented in a total 

abundance of 600 animals would be 3.0. 

H' = −Σ [(ni/N) ln(ni/N)] (3) 

where the sum is over S species. 

• Pielou’s Index (Pielou 1977) is a commonly used measure 

of community evenness. It expresses the observed value of 

Shannon’s Index relative to the maximum possible value 

(ln(S*)). Pielou’s Index, given as J' in Equation (4), varies 

between 0 and 1, and it generally co-varies with Shannon’s 

Index. 

J' = H'/ln(S*) (4) 

where H' is Shannon’s Index and S* is the number of taxa. 

Physicochemical Characterization 

of Water Column 

The water at each bridge was characterized by measuring 

temperature, pH, salinity, hardness, depth, total suspended 

solids, and total volatile solids. A point-in-time assessment of 

the environment’s ability to disperse pollutants associated 

with each bridge was assessed by measuring current speed 

and water depth at intervals across the channel, with emphasis along the designated transect line. Note that this point-intime profile has significant limitations in interpreting sediment data because sediment concentrations are the result of 

accumulations over time. Obviously, river currents and 

depths change with each season. River depth and speed are 

provided to give the reader a sense of the nature of the 

stream. To obtain an understanding of the hydrodynamics 

associated with each bridge, this information must be considered in light of the photographic evidence provided and the 

sediment grain size distribution along the transect. 

Figure 5—Conceptual use of graphical techniques and 

regression analysis to evaluate the biological response 

to treated-wood bridges. 

Cup shape indicates 

a negative biological response 

Dome shape indicates 

a positive biological response 

Straight line indicates 

no response 

(positive or negative) 

to the bridge 

Biological response 

Downstream Bridge Upstream 
10 

Physicochemical Characterization 

of Sediments 

Sediments were characterized by determining the sediment 

grain size distribution and percentage total volatile solids. 

Preservative constituents (PAH, penta, copper, chromium, 

and arsenic) were measured in the water column or sediments, or both, as appropriate at each bridge.

Characterization of Biological 

Response 

The biological response at each bridge was assessed by 

characterizing the invertebrate community collected by either 

a petite Ponar grab or a 0.0309-m2

 sampling device of proprietary design. In addition, four laboratory bioassays were 

conducted on either the water (for CCA-treated bridges) or 

sediments (for penta- and creosote-treated bridges). 

Cleaning of Sample Containers 

and Equipment 

Sample bottles were obtained, precleaned, from the laboratory subcontracted to conduct the analyses. Glass bottles 

were used in all instances. Samples for PAH and penta analyses were washed with a phosphate-free detergent solution, 

followed by thorough rinses with hot tap water and analytefree water. This was followed by an acetone rinse and a final 

rinse using high-purity methylene chloride. Lids were placed 

on the containers during the final rinse step because the 

solvent could rinse plastic from the interior screw threads 

onto the Teflon lining. 

Sample bottles for metal chemistry were cleaned in a detergent solution, rinsed with metal-free water, and soaked overnight in a covered bath containing dilute reagent-grade nitric 

acid. The bottles were then rinsed in metal-free water. 

New platinized silicone tubing was used to collect water 

samples at the CCA- and penta-treated bridges. Separate 

pieces of tubing were used at each bridge. The tubing was 

soaked overnight in a 20% nitric acid bath at room temperature, then rinsed thoroughly in distilled water. Each 7.6-mlong piece of tubing was coiled and stored in a sealed bag 

until opened for use. 

Sample Collection and Field 

Processing 

Water column samples for chemical analysis were collected 

using a Masterflex LS Sampling Pump (Cole Parmer Instrument Company, Vernon Hills, Illinois) and platinized Teflon 

tubing with an epoxy-coated lead weight. Samples were 

filtered in the field across Corning Costar Corporation Nucleopore filters (0.45 μm) (Corning, New York) and acidified. A single piece of tubing was used for each bridge. 

Samples were collected at the upstream reference station 

first, followed by the furthest downstream station, then proceeding upstream to a location under the bridge. Water samples for bioassays were collected in new 2-L bottles by holding the bottle under the water at a depth of about 0.5 m and 

unscrewing the container’s top. A total of 4 L of water was 

collected for each bioassay. Similar procedures were used to 

collect water samples in 500-mL bottles to determine total 

suspended and total volatile solids. 

Sediments were sampled using either a stainless steel petite 

Ponar grab (0.0225 m2

) or the stainless steel grab described 

in Figure 6, which samples an area of 0.0309 m2

. Samplers 

were washed with detergent in hot water, then rinsed copiously in hot water followed by a rinse in distilled water prior 

to sampling each bridge. Reference station sediment samples 

were collected first, followed by the most downstream stations, then proceeding toward the bridge. 

The sampler described in Figure 6 was used on rocky substrates and where heavy vegetation cover prevented use of 

the petite Ponar grab. Separate samples were collected for 

physicochemical analyses and infaunal analysis. The entire 

sample was processed to evaluate infauna. In vegetated areas, 

this included the vegetation growing within the footprint of 

the sampler. Similarly, this sampler was used in the rocky 

substrates associated with the penta-treated bridges surveyed 

in Washington and Oregon. Samples for physicochemical 

analyses were taken from the upper 2.0 cm of the sediment 

column. 

Sample Documentation and Handling 

Samples were tightly capped in prelabeled bottles and stored 

on frozen phase change gel packs to maintain ≤4o

C. Samples 

Figure 6—Stainless steel sampler designed for use by 

scuba divers. The sampler completely encloses a 

sediment sample covering an area of 0.0309 m2

, preventing loss of sample as it is returned to the surface. 
11 

were shipped by overnight delivery service to the appropriate 

analytical laboratory using chain of custody procedures that 

comply with requirements (ASTM 1988). 

Sediment and Water Analyses 

for Preservatives 

The following summarizes the methods used to evaluate the 

water and sediments for preservative constituents lost from 

pressure-treated wood used in constructing timber bridges. 

Analyte Method Detection limit 

Polycyclic aromatic 

 hydrocarbons (PAHs) 

HPLC (EPA 8310) See Appendix A 

Penta GC/ECD (EPA 8151) 2 μg/L or /kg

Copper (water) ICP-MS (EPA 200.10) 0.030 μg/L 

Chromium (water) GFAA (EPA 200.10) 0.075 μg/L 

Arsenic (water) GFAA (EPA 200.10) 0.310 μg/L 

Copper (sediment) ICP (EPA 6010) 0.100 mg/kg 

Chromium (sediment) ICP (EPA 6010) 0.500 mg/kg 

Arsenic (sediment) GFAA (EPA 7060) 0.200 mg/kg 

Sediment and Water Analyses 

for Conventional Parameters 

The following procedures were used to evaluate conventional 

parameters in sediments and water: 

• Total volatile solids (TVS) analysis was accomplished on 

50-g surficial sediment samples. Samples were dried at 

103 ± 2o

C in tarred aluminum boats that had been precleaned by ashing at 550o

C for 30 min. Drying was continued until no further weight reduction was observed. The 

samples were then ashed at 550o

C for 30 min. or until no 

additional weight loss was recorded in successive weighings. Total volatile solids were calculated as the difference 

(percentage) between the dried and ashed weights. 

• Sediment grain size (SGS) analysis was accomplished 

using 100-g surficial sediment taken from the top 2 cm of 

the sediment column. The sample was wet-sieved on a 

0.64-μm sieve. The fraction retained on the sieve was 

dried in a 92o

C oven and processed using the dry sieve and 

pipette method (Plumb 1981). The sieves used for the 

analysis had mesh openings of 2.0, 0.89, 0.25, and 

0.064 μm. Particles passing the 0.064-μm sieve during wet 

sieving were analyzed by sinking rates in a column of water (pipette analysis). 

• Reduction–oxidation potential discontinuity (RPD)—A 

clear, 2.0-cm-diameter corer was inserted into the 

undisturbed sediment in samples collected for physicochemical analysis. The depth at which the sediment color 

changed from the background color to black (indicative of 

the precipitation of iron sulfides associated with anaerobic 

conditions) was measured in centimeters below the sediment–water interface. 

• Benthic infaunal analysis—Invertebrate samples were 

sieved on 500-=P VWDLQOHVV VWHHO VFUHHQV XVLQJ water that 

was filtered to 37 μm. The material retained on the sieve 

was fixed in isotonic water using either 10%, boraxbuffered formaldehyde or nontoxic HistoCHOICE (Solon, 

Ohio) fixative when commercial shipment was necessary. 

Fixed samples were thoroughly washed in 10-μm-filtered 

laboratory water, following 96 h of fixation, and preserved 

in 70% ethanol. Invertebrates were picked from the background matrix under a microscope at 10× to 40× magnification. A second technician re-picked 10% of the samples. 

When more than 5% additional organisms were obtained 

during the quality assurance check, all samples picked by 

the failed technician were re-picked. 

Invertebrates were identified to the lowest practical level 

(generally genus for all orders except Chironomids, which 

were sometimes identified only to family or tribe). A reference collection containing representative samples for each 

taxon was developed and archived in 70% alcohol. Following identification, all samples were archived. The reference collection will be retained permanently. Individual 

samples will be retained for 3 years (until October 2001). 

• Current speed and water depth—As a result of the very 

slow currents at most sites, current speed was assessed by 

determining the time (seconds) required for a neutrally 

buoyant drogue to travel a distance of 2 m along a polyvinyl chloride (PVC) pipe fixture designed specifically for 

this purpose. Water depths were measured using a meter 

stick. Three replicate measurements were made at each 

sample station. 

• Water total suspended solids (TSS by APHA [n.d.] 

2540B)—A representative water sample (approximately 

300 mL) was thoroughly mixed, filtered through a 45-μm 

glass filter that had previously been ashed at 550o

C, and 

weighed. The filter was washed three times with 10 mL of 

distilled water with complete drainage between washings. 

Suction was continued for 3 min after filtration was complete. The filter was then dried at 103 ± 2o

C until no 

weight loss was recorded on successive weighings to 

0.1 mg. The difference between the filter’s original weight 

and the weight with dried residue was considered to be the 

TSS. The value was expressed as milligrams per liter. 

• Alkalinity was determined titrimetrically using APHA 

Method 403. 
12 

• Hardness was determined using the EDTA Titrimetric 

Method (APHA Method 314B.2). 

• Water total volatile solids (TVS by APHA Method 

2540 B)—The filter, containing the residue used to measure TSS, was ashed at 550o

C in a muffle furnace for 1 h 

and reweighed. The weight loss on ignition, expressed in 

milligrams per liter, was taken as the TVS. Quality assurance involved weekly calibration of the four-place balance, 

routine running of blank filters, and triplicate analyses on 

5% of the samples (minimum of one). 

• Temperature, salinity, dissolved oxygen, and pH—

Temperature was measured in the field using both a mercury thermometer and a temperature meter. A dissolved 

oxygen meter was used to measure dissolved oxygen in 

situ and a microcomputer was used to measure pH. Each 

meter was calibrated immediately before use at each bridge 

site. A two-point calibration (pH 7 and 10) was used for 

the pH meter. 

• Bioassay protocols—Four bioassays were conducted at 

each of the bridges evaluated during this study. As a result 

of differences in sediment characteristics and the marine 

environment at the new bridge site in Sandestin, Florida, 

several types of organisms were used at different bridge 

sites. Laboratory control sediments and a local upstream 

reference station at each bridge provided two levels of control in these bioassays. The results at each of these control 

levels were compared with amphipod survival at three 

treatment stations located 0.0, 1.8, and 9.9 m downstream 

from the perimeter of each bridge. Cause and effect relationships were examined using correlation and regression 

analyses to compare sediment levels of preservative constituents with amphipod survival. 

Hyalella azteca—Ten-day amphipod sediment bioassays were conducted on sediments from four stations 

at each of the penta- and creosote-treated bridges. 

Sediments were evaluated because that was the environmental compartment where maximum concentration levels of either PAH or penta were anticipated 

and observed. These tests were conducted by MEC 

Analytical Systems, Inc., in Carlsbad, California, using MEC Protocol P034.0. Methods employed in their 

program followed the general procedures outlined by 

ASTM (1997). 

Menidia berylina—A 96-h acute bioassay was conducted on water from the newly constructed CCA 

timber bridge in Sandestin, Florida. Water, rather than 

sediment, was tested at both CCA bridges included in 

this study because the test species are most sensitive to 

the dissolved cupric ion (Cu+2). This particular test 

animal was chosen because this was a marine site with 

salinity measured at 24.9 ppt (parts per thousand). 

MEC Testing Protocol No. P009.1 was used at 

25.5o

C. Copper sulfate (CuSO4) was used as a reference toxicant with a 96-h LC50 of 114.0 μg/L. 

Daphnia magna—A 96-h acute bioassay was conducted on the water from the old CCA timber bridge 

evaluated in Sandestin, Florida. The protocol given in 

EPA 600/4-90/027 (4th edition) was used at a test 

temperature of 25o

C. Five ostracods were included in 

each of four replicated test chambers (20 organisms 

total). Sodium chloride was used as a reference 

toxicant. 

Quality Assurance Requirements 

and Data Qualifying Criteria 

The following are quality assurance requirements for the 

analysis of copper, chromium, arsenic, penta, and PAHs 

sampled during the timber bridge environmental risk assessment: 

Requirement Data qualifier criteria 

Method blank 

(1 per 20 samples) 

No analyte detected in blank 

Replicate analyses 

(1 per 20 samples) 

≤20% RPD 

Matrix spike 

(1 per 20 samples) 

75% to 125% spike recovery 

Surrogate recovery 

(all samples) 

±95% confidence interval 

Sediment Grain Size and Total Volatile Solids 

Sediment grain size (SGS) and total volatile solids (TVS) 

require triplicate analyses on one or a minimum of 5% of the 

samples. The root square deviation (RSD) should be <20% 

for same sample results. 

Infaunal Analysis 

Infaunal analysis standard AES QA/QC procedures were 

used (PSEP 1986). 

Results and Discussion 

Creosote-Treated Bridges 

Since 1865, creosote has been widely used in the United 

States to protect wood from attack by fungi, marine borers, 

and insects. Creosote is a distillate derived from coal tar 

produced by the carbonization of coal and consists primarily 

of liquid and solid aromatic hydrocarbons. 

The treatment of wood products is generally accomplished in 

accordance with the American Wood-Preservers’ Association 
13 

(AWPA) C1–92 (AWPA 1992). Average creosote penetration using the “empty cell” process in Douglas-fir is 38.1 mm 

(1.5 in.), and retention is 320 kg/m3

 (20 lb/ft3) for marine 

applications and 272 kg/m3

 (17 lb/ft3) for freshwater 

(Arsenault 1992, Baechler and Alpen 1962). In 1996, AWPA 

reduced the required retention in temperate marine waters 

from 320 to 256 kg/m3

 (20 to 16 lb/ft3). 

Creosote is a complex mixture of at least 160 detectable 

hydrocarbon compounds; all 18 major components are cyclic 

and aromatic. According to Environment Canada (1992), 

80% of creosote is composed of PAHs. Some low-molecularweight creosote compounds, such as the naphthalenes, have a 

density less than one. However, most components are heavier 

than seawater and sink in a water column. 

The public has voiced concern because creosote contains 

many of the 16 PAHs known to be acutely and chronically 

toxic to marine animals. Several of these compounds, most 

notably benzo(a)pyrene, can degrade to carcinogenic, teratogenic, and mutagenic intermediates during metabolism. As 

with most potentially harmful substances, pathological responses are a matter of exposure (concentration and length of 

time). 

Sources of PAH 

Polycyclic aromatic hydrocarbons (PAHs) are formed by a 

variety of processes, including indirect and direct biosynthesis, fossil fuel production and distribution, and incomplete 

combustion of organic matter. When formed, PAHs can be 

transported into an aquatic environment through several 

pathways, including fossil fuel distribution, storm water 

runoff, sewage effluent, and atmospheric deposition. 

• Biosynthesis—Indirect biosynthesis of PAH occurs when 

extended quinones and related polycyclic materials (mostly 

plant and animal pigments) are exposed to the reducing 

conditions found in anoxic sediments. The resulting PAHs 

tend to accumulate in the sediments where they were 

formed. De novo biosynthesis of PAHs by aerobic and anaerobic bacteria, fungi, and plants is controversial. However, Mallet and others (1972) concluded that both aerobic 

and anaerobic bacteria can biosynthesize benzo(a)pyrene 

(B(a)P) and certain other PAHs using fatty acids, sterols, 

plant pigments, and aliphatic terpenes as substrates. In 

most cases where PAH biosynthesis has been reported, accumulation of PAH in the organisms purported to have 

synthesized them could also have been attributed to uptake 

of PAH from exogenous sources. In light of the literature 

reviewed, it appears that PAH biosynthesis may occur to a 

limited extent under special environmental conditions 

when necessary bacterial growth substrates are present. 

Eisler (1987) suggested that, on a global scale, biosynthesis annually contributes about 2.7 × 106

 kg (6 × 106 lb) of 

PAH to aquatic environments (Table 3). 

• Fossil fuels, including peat, coal, and petroleum, are relatively rich in complex assemblages of PAHs. These compounds reach aquatic environments through surface runoff, 

in wastewater, and as a result of petroleum spillage. Eisler 

(1987) estimates that spilled petroleum contributes 168 × 

106

 kg (374 × 106 lb) of PAH to aquatic environments each 

year (Table 3). This source overwhelms all others in terms 

of global inputs.

• Pyrolysis of organic matter between 400oC and 2000oC 

results in the generation of a wide variety of PAHs. Reducing conditions (insufficient oxygen) in the pyrolytic environment favor PAH production. Forest and grass fires, industrial processes, heating, power generation, and 

petroleum refining release significant amounts of PAH into 

the atmosphere. These products of combustion are subject 

to chemical- and photo-oxidation. However, their residence time in the atmosphere is long enough to allow wide 

dispersal, and they are a major source of PAHs in aquatic 

environments. According to Eisler (1987), forest and prairie fires and agricultural burning annually release nearly 

32.4 × 106

 kg (72 ×106 lb) of PAHs into the atmosphere. 

This is three times the amount from all other pyrolytic 

sources combined. 

• Petrolytic sources—Johnston and Harrison (1984) reported 

that B(a)P deposition along a United Kingdom 

motorway is 2.8 μg/m2

-week. Note that B(a)P is approximately 0.5% to 2.5% of some PAH mixtures, and a direct 

extrapolation suggests that the total PAH loading along a 

well-used highway may be 560 μg/m2

-week. Winter levels 

of PAH in coastal areas are greater than summer levels. 

This is attributed to increased pyrolytic input from the 

burning of fossil fuels for power generation and heating 

(Bouloubassi and Saliot 1991). Broman and others (1990) 

suggested that primary PAH inputs in the Baltic region 

were from exhaust emissions associated with automobiles, 

domestic heating, refuse incineration plants, ships, and aircraft. Neff (1979) reported that little-used (224 km) motor 

oil contained 6.4 ppb (parts per billion) B(a)P, equivalent 

to nearly 1,280 ppb total PAH. (In this paper, billion is 

equivalent to 109

.) Dunn and Stich (1976) found up to 

Table 3—Global sources of PAHs to aquatic 

environmentsa 

Source PAH (kg/year (lb/year)) 

Petroleum spillage 168,620,000 (374,782,200) 

Atmospheric deposition 

(from combustion) 

49,603,500 (110,230,000) 

Wastewater 4,365,108 (9,700,240) 

Surface land runoff 2,916,686 (6,481,524) 

Biosynthesis 2,678,589 (5,952,420) 

a

Eisler 1987. 
14 

22,000 ppb B(a)P in well-used crankcase oil. This is 

equivalent to 4.4 ppt total PAH. In 1989, Washington state 

had 4,179,000 vehicles registered. These vehicles produced 63,000 m3

 (21,000,000 gal) of used crankcase oil 

each year. That may represent as much as 281,250 kg 

(625,000 lb) of PAHs available to the environment. 

• Industrial and domestic wastewaters are rich in PAHs. 

Secondary sewage treatment removes some PAHs, but 

most are released to aquatic environments through sewage 

treatment plant outfalls. Eisler (1987) notes that untreated, 

raw sewage contains 100 to 500 ppb total PAHs and sewage sludge contains 200 to 1,750 ppb PAH. Goyette and 

Boyd (1989) recorded sediment PAH concentration levels 

of 17 μg TPAH/g dry sediment near a major combined 

sewage–storm water outfall discharging into Vancouver 

Harbor. Sediment PAH levels in the central areas of the 

harbor were consistently in the range of 2 to 5 μg PAH/g 

dry sediment. Hoffman and others (1984) noted that storm 

water runoff from urban areas and highways accounted for 

71% of the high-molecular-weight PAHs and 36% of the 

total PAH loading to Narragansett Bay. More than 30% of 

all pyrolytic PAHs in the coastal sediments of Washington 

State are supplied by riverine transport of suspended particulate materials, and direct atmospheric input accounts 

for a maximum of 10% (Prahl and others 1984). 

Observed Levels of PAH 

As a result of the many natural sources, PAHs have been 

present in aquatic environments for thousands of years. 

However, significantly increased levels have been recorded 

in ocean sediments since the turn of the 20th century. Neff 

(1979) reviewed the distribution of PAHs in aquatic environments. He found very low (<1 to 2 ppb) levels of PAHs in 

the water column of pristine areas and recorded low sediment 

contamination in pristine areas (< 0.050 μg/g) but higher 

levels (to 15 μg/g) associated with industrialized areas or 

human population centers. Eisler (1987) found a similar 

distribution of sediment PAH levels. Levels in pristine areas 

of Alaska, Africa, and the Amazon Basin ranged from 0.005 

to 0.544 μg/g. Levels in urbanized and industrialized areas 

ran as high as 791 μg/g in the United Kingdom. Cerniglia and 

Heitkamp (1991) measured sediment PAH levels ranging 

from 0.005 μg/g for an undeveloped area in Alaska to 

1,790 μg/g at an oil refinery outfall in Southampton, 

England. Sediment PAH concentration levels in other industrialized areas ranged from 0.198 to 232 μg/g. Transportation 

systems represent a significant source of PAHs associated 

with internal combustion engines, lubricants, herbicides, 

and asphalt and tars used in road surfacing. The following 

summarizes Eisler’s (1987) assessment of PAHs loading 

to aquatic environments from the most common sources: 

There is a consistent thread running through research and 

reviews by Bouloubassi and Saliot (1991), Neff (1979), and 

Eisler (1987). These authors indicate that PAHs are present 

in aquatic and terrestrial environments. There are numerous 

natural sources of PAH, including volcanoes, forest and 

prairie fires, natural oil seeps, and biosynthesis. According to 

the sediment record, PAHs have been present in our environment since there was life. Historically, the most natural 

sources of PAH were either the result of biosynthesis, in 

which case the PAH remained stable in the anoxic environments where they were formed, or the inputs resulted from 

volcanoes or forest and prairie fires. These pyrogenic PAHs 

were widely distributed over large geographic areas and 

resulted in relatively low background levels (< 0.01 to 

0.05 μg/g). 

Anthropogenic inputs from oil spills, wastewater, storm 

water, and petrolysis tend to be concentrated in urban and 

industrial areas. This may result in concentrating PAHs in 

aquatic environments and the sediments that underlie them. It 

is these high concentration levels that are of concern. Petroleum spills represent the major PAH input to our oceans and 

can have devastating, although probably short-term (2 to 

5 years), impacts on aquatic resources. 

The literature suggests that sediment PAH levels of less than 

0.2 to 0.5 mg/kg may be expected in pristine areas. Worldwide, urban areas have much higher background sediment 

PAH levels in the range of 0.5 to several milligrams PAH per 

kilogram of dry soil or sediment. Heavily industrialized areas 

may have PAH levels of 10 to several hundred milligrams 

PAH per kilogram.

Fate of PAH 

PAHs form a family of compounds, and the routes of degradation and fates are different for the major classes of PAH. In 

water, PAHs evaporate, disperse into the water column, 

become incorporated in bottom sediments, concentrate in 

aquatic biota, or experience oxidation and biodegradation. 

The most important degradative processes for PAHs in the 

marine environment are photo-oxidation, chemical oxidation, 

 PAH loading to aquatic 

environment 

Source kg/PAH/year lb/PAH/year 

Petroleum spillage 374,782,200 

Atmospheric deposition 

(from combustion) 

 110,230,000 

Wastewater 4,365,108 9,700,240 

Surface land runoff 2,916,686 6,481,524 

Biosynthesis 2,678,589 5,952,420 
15 

and biological transformation by bacteria and animals (Neff 

1979). Most PAHs in aquatic environments are associated 

with particulate materials and only about a third are present 

in dissolved form. Dissolved PAHs will likely degrade rapidly through photo-oxidation (USEPA 1980b). They degrade 

most rapidly at high concentration levels, at elevated temperatures and oxygen levels, and at high levels of solar irradiation. Different PAHs vary significantly in their relative 

sensitivity to photo-oxidation. 

Because of their low aqueous solubility and hydrophobic 

character, the higher molecular weight PAHs readily adsorb 

to particulate materials and solid surfaces in water. The 

ultimate fate of PAHs that accumulate in sediments is believed to be biotransformation and degradation by benthic 

organisms (USEPA 1980b). High-molecular-weight PAHs 

degrade slowly in anaerobic sediments (Neff 1979). 

Cerniglia and Heitkamp (1991) discussed microbial degradation of PAHs in aquatic environments. They noted that a 

wide variety of bacteria, fungi, and algae have demonstrated 

ability to metabolize PAHs. Low-molecular-weight PAHs, 

such as naphthalene, degrade rapidly. The higher molecular 

weight PAHs, such as benz(a)anthracene and 

benzo(a)pyrene, are more resistant to microbial attack. 

Cerniglia and Heitkamp (1991) also note that the most rapid 

biodegradation of PAHs occurs at the water–sediment interface. This is because prokaryotes oxidize PAH as a first step 

in metabolism. Deeper sediments usually contain little oxygen, thus inhibiting microbial metabolism. 

Cerniglia and Heitkamp (1991) summarized the available 

literature describing the half-life of PAHs in various environments. The results are highly variable and depend on 

PAH species together with a range of environmental and 

biological factors. Bacterial communities in polluted areas 

metabolize PAHs more quickly than do communities in 

unpolluted areas. Lighter weight PAHs are metabolized more 

quickly than heavier PAHs. Naphthalene has a short turnover 

time (days to weeks), whereas the five-ringed benzo(a)pyrene 

has a long turnover time (years under unfavorable conditions). 

Ingram and others (1982) observed that the concentration of 

creosote in leaching vats increased to greater than 700 ppb in 

the first 72 h, then decreased to less than 34 ppb at the end of 

20 days. They attributed that decrease to bacterial metabolism of the low-molecular-weight PAHs being leached from 

the pile sections in their study. 

Tagatz and others (1983) noted that creosote concentration 

levels decreased by 42% during 8 weeks in sediments artificially contaminated as part of their mesocosm studies. They 

attributed the decrease to microbial metabolism. 

Neff (1979) attempted to integrate the degradative processes 

associated with PAH removal from aquatic environments. 

He concluded that the residence time of PAH in water is 

brief. The lower molecular weight aromatics (benzene to 

phenanthrene) are removed primarily by evaporation and 

microbial activity. Higher molecular weight PAHs are removed mainly by sedimentation and photo-oxidation. Degradation of PAH by animals in the water column is of minor 

importance. In nutrient-rich, biologically active, aerobic 

sediments, the degradation of PAHs is dramatically increased 

by healthy bacterial and fungal communities. In anaerobic 

sediments, the higher molecular weight PAHs (4 to 7 rings) 

may persist for years. 

Bioconcentration, Bioaccumulation, 

and Biomagnification 

Bioconcentration and bioaccumulation of contaminants are 

of special importance because some aquatic species, most 

notably bivalves, have demonstrated an ability to rapidly 

bioconcentrate some contaminants in water to high tissue 

levels. The concern is that persistent contaminants may move 

up the food chain, biomagnifying to higher concentration 

levels in each trophic level, until contaminants found at 

nontoxic levels in the ambient environment reach levels 

where they do cause stress and disease. For effective biomagnification and movement of contaminants through the 

food chain, several conditions must be met. First, organisms 

must have the ability to bioconcentrate low levels of contaminants from the water column or bioaccumulate PAH 

from sediments or their food. Second, these contaminants, or 

their toxic metabolic intermediates, must be retained, unaltered, in the tissues of the organism until it falls prey to an 

animal at a higher trophic level. 

Several factors mitigate biomagnification. If contaminants 

are not absorbed, they cannot accumulate. Numerous organisms, particularly vertebrates, have the ability either to metabolize or excrete organic contaminants. The gut, liver, 

kidney, and gall bladder are common sites of PAH concentration, metabolism, and excretion in vertebrates. If the 

contaminants are either rapidly excreted or metabolized to 

nontoxic compounds, then the chain is broken and biomagnification is not effective in passing contaminants upward 

through the food chain. DDT is an excellent example of a 

persistent compound that is bioconcentrated from low levels 

in the water to higher levels; first in plankton, then in fish, 

finally in bird populations with devastating consequences. 

Neff (1982) reported that most aquatic organisms bioconcentrate PAHs from low concentration levels in the ambient 

water to higher tissue levels. Bioconcentration factors 

(BCFs) are predicted by the octanol–water partition coefficients (Kow) associated with individual PAH compounds. 

Bivalve mollusks, particularly the commercially important 

mussel (Mytilus edulis) and oysters of the genera Ostrea and 

Crassostrea, have received far more attention than other 

aquatic invertebrates, plants, or fish. Bivalve mollusks are 
16 

excellent subjects for monitoring pollutants because they 

filter substantial quantities of water over large and highly 

permeable gills. For these reasons, mussels have been the 

subject of numerous studies. Many of these studies have 

focused on the accumulation of metals and the carcinogenic 

molecule benzo(a)pyrene. 

Benzo(a)pyrene levels recorded by Neff (1979) for uncontaminated areas fall in the undetectable to perhaps 50 μg/L 

range. Dunn and Stich (1976) recorded tissue levels averaging 59 μg/g in mussels from areas associated with marinas 

and higher levels, averaging 402 μg/g, in mussels growing on 

creosote-treated pilings. Dobroski and Epifanio (1980) found 

that direct uptake of B(a)P from seawater by diatoms was 

much greater than the rate of trophic transfer from the diatoms to clam larvae. 

Eisler (1987) recorded elevated PAH concentrations, especially benzo(a)anthracene, chrysene, fluorene, phenanthrene, 

and pyrene, in oyster tissues and sediments from the vicinity 

of marinas. These levels are notably higher in cooler months, 

when lipids and glycogen are being stored preparatory to 

spawning (Marcus and Stokes 1985). 

For mussels, the general trend towards lower levels of higher 

molecular weight PAHs relative to the levels in associated 

sediment suggests an uptake mechanism that involves the 

solution of PAH in water. Supporting this hypothesis is the 

observed rapid turnover and shorter half-life of the more 

soluble, lower molecular weight PAHs (Dunn 1980, Eisler 

1987). This suggests that the more soluble (and more 

bioavailable) PAHs are effectively removed from sediments 

and metabolized by bivalves. The higher molecular weight 

PAHs (associated with chronic stress and genetic disorders) 

remain in the sediments because of their low solubility. 

However, when absorbed, PAHs are more slowly metabolized by bivalves. 

PAH levels in fish are usually low because this group rapidly 

metabolizes all PAHs (Lawrence and Weber 1984, West and 

others 1986a,b) or excretes them. High concentration levels 

of PAH are typically found in the gut, liver, and bile. Raw 

fish from unpolluted or moderately polluted water seldom 

contains detectable amounts of PAHs. However, smoking 

and cooking of fish can increase PAH content to significant 

levels. 

Neff (1982) reported BCFs for several PAHs in the clam 

Rangia cuneata. Note that the BCFs, which range from 6.1 to 

32, are for PAHs dissolved in water. Eisler (1987) summarized BCF values from the literature. The BCFs reported in 

his paper contradict his assertion that bivalves accumulate 

PAHs more rapidly than fish. For all the values given in his 

review, the average BCFs are 82 for bivalves (n = 8) and 

6,844 for fish (n = 34). Note that Eisler’s (1987) paper reported bioconcentration values from 6 to 236 in the clam, 

Rangia cuneata. Four of the five values were less than 33. 

For fish, bioconcentration values range from 44 to 82,916, 

with most values in the 100 to 1,000 range. 

The ultimate fate of most high-molecular-weight PAHs 

deposited in aquatic environments is sedimentation. Roesijadi and others (1978) examined the accumulation of crude 

oil in Prudhoe Bay and specific PAHs from oil-contaminated 

sediments by three infaunal invertebrate species, the sipunculid worm Phascolosoma agassizii and the clams Macoma 

inquinata and Protothaca staminea. They found that efficiency of PAH uptake from sediments was much lower than 

from water. Bioconcentration factors for uptake of the four 

PAHs from contaminated sediments were 0.2 or less, indicating no significant bioconcentration of PAH by this route. 

However, BCFs for uptake of these four PAHs from seawater 

were in the 10.3 to 1,349 range, indicating a low to moderate 

potential for bioconcentration. 

Eisler (1987) suggested that bivalves readily take up PAHs 

from sediments. This hypothesis is contradicted by the results 

of numerous studies. O’Connor (1991) found that at 117 

National Status and Trend Sites where there were both mollusks and fine-grained sediments, the average ratio of mollusk tissue to sediment concentration was only 1.2 for total 

PAHs. He also noted that mollusks accumulate the lowmolecular-weight (and more highly soluble) PAHs to a 

greater extent (2.0) than the high-molecular-weight PAHs 

(0.64). Eaton and Zitko (1978) noted that PAH levels in 

clams and mussels were two orders of magnitude below those 

detected in sediments. Neff (1979) cites Perdriau’s (1964) 

finding that in no case did benthic animals contain elevated 

levels of B(a)P when compared with sediment concentration 

levels. Tissue concentrations in the animals were, on average, 

36% of the sediment concentration. 

The previously cited studies and Neff (1982) lead to the 

general conclusion that sediment-adsorbed PAHs are not 

readily assimilated by benthic animals. Accumulation of 

PAHs from sediment, when it occurs at all, may be attributed 

in large part to uptake of PAHs desorbed from sediment 

particles into the interstitial water. This hypothesis is supported by Swartz and others (1989) who concluded that the 

concentration of chemicals in interstitial water is the primary 

determinant of sediment toxicity, not the bulk concentration 

in the sediment. 

Depuration of PAH 

Southworth and others (1978) found a half-life of less than 

1 h for all PAHs metabolized by Daphnia pulex. Jackim and 

Lake (1978) reported that the half-life of PAHs in most 

bivalves is about 2 to 16 days. These studies suggest that 

PAHs are either rapidly metabolized or excreted, at least by 

these species. 
17 

Biomagnification of PAH in the Food Chain 

Neff (1979) reported that the annelid Neanthes arenaceodentata had little, if any, ability to accumulate 2-methylnaphthalene from its food. However, the situation is quite 

different in marine crustaceans and fish, where uptake from 

food is much more efficient than uptake from water. Arthropods (for example, crabs, amphipods, shrimp) rapidly accumulate the lighter weight PAHs and very rapidly excrete or 

metabolize these compounds. The half-life of B(a)P in 

Callinectes sapidus is 6 days. Neff (1979) concluded that all 

results dramatically demonstrated the importance of metabolism in eliminating PAHs from contaminated crustaceans. 

Broman and others (1990) examined the trophic transfer of 

PAHs in a study involving seston, the blue mussel (Mytilus 

edulis), and the eider duck (Somateria mollissima). Contrary 

to biomagnification, they observed decreasing PAH concentration levels with increasing trophic levels. 

Bioaccumulation Summary

Aquatic organisms are able to efficiently bioconcentrate 

PAHs from the water column. It appears that direct transfer 

from sediments to organisms living within and on those 

sediments is minimal. Benthic organisms rarely contain 

higher concentration levels of PAH than are found in the 

sediments in which they live. PAHs are rapidly metabolized 

and excreted by vertebrates and arthropods. In bivalves, 

which do not efficiently metabolize PAHs, the half-life of 

most PAHs examined was 2 to 16 days. These data suggest 

that PAHs are not persistent in the tissues of aquatic species 

and that movement of PAHs through food chains to higher 

trophic levels is minimal, if it occurs at all. 

Neff concluded that 

From the limited data available, it would appear that 

there are large interspecific differences in ability to absorb and assimilate PAH from food. Polychaete worms 

have a very limited ability to absorb and assimilate 

PAH, whereas fish absorption of PAH from the gut is 

limited and variable depending on species of fish, the 

PAH, and possibly the food matrix in which PAH is 

administered. Crustaceans, on the other hand, apparently readily assimilate PAH from contaminated food. 

In all cases where assimilation of ingested PAH was 

demonstrated, metabolism and excretion of PAH were 

rapid. Thus, the potential for food chain biomagnification of PAH seems to be limited. For such biomagnification to occur, the material must be readily absorbed 

from food, and once assimilated, it must be relatively 

resistant to metabolism or excretion. (Neff 1979) 

Creosote and PAH Toxicity 

Aquatic organisms have been exposed to background levels 

of PAH for eons. The diversity of life in aquatic environments attests to the ability of aquatic species to tolerate these 

background levels (1 to 2 μg/L in the water column and 

0.010 to 0.50 μg/g in sediments). At what level do PAHs 

cause pathological responses at the organismal and population levels? In answering that question, we need to consider 

two types of toxicity: acute and chronic. 

Acute toxicity causes observable physiological lesions and is 

usually measured by mortality. PAH can interact with cells in 

several ways to cause toxic responses. For example, PAHs 

may bind reversibly to lipophilic sites in the cell, thereby 

interfering with cellular processes. Potentially impacted and 

important intracellular organelles include lysosomes, which 

contain strong enzymes important in intracellular digestion 

of complex organic molecules and in the immune response. 

Increased lysosomal membrane permeability can result in the 

unregulated flow of these enzymes into the cytoplasm or 

blood serum with pathological consequences including autophagy. Eisler (1987) noted that the lower molecular weight, 

unsubstituted PAH compounds, containing two or three 

rings, such as naphthalene, fluorene, phenanthrene, and 

anthracene, have significant acute toxicity to some organisms 

whereas the higher molecular weight, four- to seven-ring 

aromatics, do not. However, these heavier molecules 

contain numerous potentially carcinogenic and mutagenic 

intermediates. 

A common measure of acute toxicity is the concentration of a 

toxicant that causes 50% mortality in a test population within 

some specified period (often 96 h). This parameter is referred 

to as the 96-h LC50. Borthwick and Patrick (1982) and Neff 

(1979) reported 96-h LC50 values for several marine animals. 

These are summarized in Table 4. 

Interestingly, in Neff’s (1979) discussion of the effects of 

PAH on aquatic animals, he cites Caldwell and others (1977) 

finding that continuous exposure to dissolved naphthalene 

Table 4—Acute toxicity of various PAHs to marine 

organisms

Species 96-h LC50 (μg/L)

Mysids (Mysidopsis bahia)

a 18 to 21 

Oysters (Crassostrea virginica)a 700 

Pink shrimp (Penaeus duorarum)

a 240 

Sheepshead minnows (Cyprinodon 

variegatus)

a 3,500 

Mosquito fish (Gambusia affinis)

a

 150,000 naphthalene 

Mosquito fish (Gambusia affinis)

b

 1,180,000 toluene 

Dungeness crab larvae (Cancer 

magister)

b 8 naphthalene 

Dungeness crab larvae (Cancer 

magister)

b 170 naphthalene 

a

Borthwick and Patrick 1982. 

b

Neff 1979. 
18 

concentrations of 19 to 170 μg/L had no effect on the survival of Dungeness crab larvae. No explanation was given for 

the low (8 ppb) value reported in Neff’s (1979) paper or the 

differences in the values reported. You might expect that 

exogenous factors contributed to the discrepancy. The LC50

values reported in the literature for most organisms and PAH 

compounds are in the 500 to 5,000 ppb range. Neff (1979) 

found that in all but a few cases the concentration levels of 

aromatic hydrocarbons that are acutely toxic to aquatic animals are several orders of magnitude greater than concentration levels found even in the most heavily polluted marine 

and fresh waters. However, sediments from polluted regions 

may contain aromatic hydrocarbons at concentration levels 

similar to or greater than those that are acutely toxic. The 

limited bioavailability of sediment-adsorbed PAHs undoubtedly renders them substantially less acutely toxic than dissolved PAH. He also noted that PAH-induced stress is cumulative and exacerbated by exogenous stress factors such as 

abnormal thermal and osmotic conditions. 

PAH Toxicity to Aquatic Plants 

The effects of various PAHs on aquatic plant growth are 

highly variable. At low concentrations (10 to 20 ppb), several 

PAHs act as a stimulant to plant growth. At 300 ppb, chrysene was observed by Boney (Neff 1979) to induce a 58% 

increase in the growth of the red alga Antithamnion plumula. 

Other PAHs (anthracene and 2-methylanthracene) caused 

declines of –20% and −12% in the same alga at 300 ppb. In 

general, PAH concentration levels greater than 1,000 ppb 

inhibit algal growth. 

Chronic Toxicity 

Neff (1979) addressed chronic stress associated with 

PAH contamination. He noted that the copepod Eurytemora 

affinis suffered statistically significant reductions in length 

of life, total number of nauplii produced, and brood size 

when exposed to 10 ppb naphthalene, 2-methylnaphthalene, 

2,6-dimethylnaphthalene, or 2,3,5-trimethylnaphthalene for 

the duration of their lives. The following discusses documented instances of chronic stress, by effect. 

Nearly all PAHs are hydrophobic and lipophilic. Thus, there 

is a potential for these compounds to become associated with 

stable lipid pools in aquatic organisms. Energy is generally 

stored as glycogen in bivalves until gametogenesis, when the 

glycogen and lipid stores are converted into eggs and sperm. 

The eggs contain significant lipid reserves and could become 

a repository for lipophilic PAH. Moore and others (1989) 

cited Lowe and Pipe’s (unpublished) observation that longterm exposure to diesel oil at 30 and 130 ppm caused a 

decrease in the mass of gametes produced by Mytilus edulis

and Macoma balthica. 

Mollusks exhibit reduced ventilation (feeding) rates at PAH 

levels as low as 30 to 40 ppb in seawater (Moore and others 

1989). The feeding inhibition probably resulted from the 

narcotic effect of hydrocarbons, particularly aromatic hydrocarbons. These compounds have a direct effect on cilia, 

muscles, and/or the nervous system, which control the mollusk’s activity. Reduced feeding rates result in a reduction in 

“scope for growth,” a commonly measured parameter that 

quantitatively describes the energy available for tissue 

growth, reproduction, and activity. In bivalves, the major 

problem caused by reduced scope for growth is poor reproductive capacity. Although this does not have immediate 

consequences at the organismal level, the long-term consequences of reduced recruitment could be significant for the 

population. 

Neff (1979) concluded his discussion of PAH-induced 

chronic toxicity by suggesting that although environmentally 

realistic PAH water column concentrations of 1 to 50 ppb 

can cause potentially detrimental, sublethal responses in 

aquatic organisms, in most cases, the PAH concentrations 

required to elicit significant sublethal responses are greater 

than those normally encountered in all but the most heavily 

polluted aquatic environments. This statement was strongly 

supported by the low levels of creosote-derived PAH 

(0.030 μg/L) observed by Goyette and Brooks (1999) in the 

immediate vicinity (15 cm) of a major creosote structure in 

Sooke Basin. 

From the preceding discussion on the uptake of PAH from 

water, food, and sediments, it appears that PAH concentrations in the water column (including interstitial water in 

sediments) are the parameters of greatest significance in 

defining chronic stress. Furthermore, it appears that sustained 

water column concentrations of 30 to 50 ppb PAH can have 

subtle, but important, chronic impacts on populations of 

marine organisms. 

Neoplasia Associated With PAHs 

Hyperplastic, preneoplastic, and neoplastic lesions have been 

reported in fish for a number of years. These same types of 

lesions are far less common in bivalves and other 

invertebrates. 

In vertebrates, enzymes produced by the cytochromium 

P-450, mixed-function oxidase (MFO), and aryl hydrocarbon 

hydroxylase (AHH) systems are responsible for initiating 

catabolism of lipophilic compounds (including PAH). These 

systems render hydrophobic molecules more water soluble 

and therefore increase their potential for excretion and detoxification. In the case of certain high-molecular-weight 

PAHs, the intermediate metabolic products of these enzyme 

systems can be highly toxic, mutagenic, or carcinogenic. 

Oxidative metabolism of some PAHs (like B(a)P) results in 

the production of arene oxides, some of which bind covalently to DNA and RNA (particularly with guanine). The 

resulting chromosomal lesions can result in unregulated cell 

growth and division (cancer). 
19 

The ability to metabolize high-molecular-weight PAHs varies 

significantly between the phyla. Among invertebrates, the 

mollusks have low AHH activity and a limited ability to 

metabolize high-molecular-weight PAHs. Arthropods and 

annelids show increased activity and some marine crustaceans have demonstrated significant cytochromium P-450, 

MFO, and AHH activity. 

Vertebrates, including fish, demonstrate high MFO, AHH, 

and cytochromium P-450 capabilities (Varanasi 1989). The 

liver is the primary site of MFO activity in fish, and the liver, 

gut, and gall bladder are primary sites of PAH concentration, 

metabolism, and excretion. Humans do not normally consume these organs. In Crustaceans, the hepato-pancreas, 

green gland (excretory organ), pyloric stomach, gills, testes, 

and eyestalks are major sites of PAH accumulation and AHH 

enzyme activity. Again, humans do not normally consume 

these tissues, although the hepato-pancreas is sometimes 

eaten as “crab butter.” 

Melanomacrophage centers are an integral part of the teleost 

immune system. Payne and Fancey (1989) observed that the 

numbers of melanomacrophage centers increased in the liver 

of fish exposed to total PAH concentrations in the range of 

25,000 to 50,000 ppb. These concentrations are found only 

in heavily polluted harbors, industrially polluted sites, or oil 

spills. Payne and others (1988) observed changes in MFO 

enzyme levels and liver fat content in fish exposed to low 

dissolved hydrocarbon levels of 1,000 ppb (perhaps even as 

low as 200 to 300 ppb PAH). 

The increased levels of P-450, MFO, and AHH enzymes in 

fish and crustaceans exposed to very high levels of PAH 

suggest active catabolism of these molecules. Enzyme induction is not necessarily a sign of stress. However, there is 

concern because some intermediate products of highmolecular-weight PAH catabolism are carcinogenic, 

mutagenic, and teratogenic. 

Bioindicator Studies 

There is growing interest in enzyme induction and genotoxicity tests as indicators of environmental risk. However, are 

genotoxicity and enzyme induction tests appropriate indicators of environmental risk? It is important to understand what 

these tests actually tell us. Effects at the organismal level, 

associated with external factors, are mediated by numerous 

levels of protection. Detrimental factors (for example, abnormal temperature, xenobiotics, desiccation, disease organisms, high levels of pollution, UV radiation) are often 

avoided by mobile animals. Sessile animals (including many 

bivalves) isolate themselves within tightly closed valves in an 

attempt to avoid harmful conditions. 

At the next level of protection, an animal’s integument isolates internal organs and structures from harmful conditions. 

The skin and gut epithelia are capable of selective absorption 

of material. For instance, high-molecular-weight PAHs, 

adsorbed to sediments, apparently pass through the digestive 

tract of many annelids without being absorbed through the 

gut epithelia. 

After foreign materials are absorbed into the blood serum 

through the skin, gills, or gut, organisms respond by sequestering them in vacuoles, metabolizing them in the liver or 

cleansing them from the serum as it passes through the 

kidney. Whether or not a molecule is metabolized or excreted 

depends largely on its ability to penetrate cell membranes. 

The plasma lemma is highly permeable to essential molecules such as glucose, amino acids, and lipids. These phospholipid bilayers are not very permeable to ions or large 

charged polar molecules. The four- to seven-ring highmolecular-weight PAHs are generally not charged and therefore do pass across the cell membrane and are actively metabolized by vertebrates. 

It is well documented that some metabolic intermediates of 

high-molecular-weight PAHs, particularly arene oxides, can 

bind covalently to guanine, producing DNA lesions, which 

may result in unregulated cell growth (cancer). These metabolic intermediates are frequently found in the digestive 

gland (liver or hepatopancreas) where metabolism is most 

active. The literature contains many citations regarding 

hepatic lesions (including hepatic carcinomas) in demersal 

fish associated with PAH-contaminated sediments. However, 

the levels of contamination observed at Eagle Harbor, the 

Duwamish River, and Elizabeth River, at which significant 

increases in hepatic carcinomas were observed, are generally 

greater than 25 to 50 mg/kg. In some cases, Eagle Harbor 

sediments contained as much as 6,000 mg/kg PAH. 

Mixed function oxidases (MFO), cytochromium P-450, 

ethoxy resofurin-o-deethylase (EROD), and aryl hydrocarbon 

hydroxylases (AHH) are important enzyme systems for the 

metabolism of high-molecular-weight PAHs. There are numerous reports in the literature suggesting that PAHmetabolizing enzyme systems are activated at sediment PAH 

levels as low as 1.0 ppm (Johnson and others 1994) 

As previously stated, intermediate PAH metabolites, such as 

arene oxides, can covalently bind to DNA, resulting in lesions. However, DNA contains numerous mechanisms that 

repair miscoded or damaged sequences. This repair is 

achieved by a suite of enzymes capable of recognizing damaged or mismatched base pairs and excising them. Environmental or random damage to DNA is not unusual, and the 

presence of nicks or double-stranded breaks in nuclear (or 

ribosomal) DNA does not often lead to unregulated cell 

growth. Increased DNA damage obviously increases the risk 

for failure of these repair mechanisms, resulting in a number 

of diseases. 

The point is that numerous levels of protection are involved 

in maintaining the biological integrity of an organism. 
20 

In evaluating environmental risks, we must recognize the 

importance of these cellular safeguards. The questions we 

ask must recognize that different levels of biological organization will respond differently to the same level of insult. 

Therefore, our questions must be posed carefully, and caution should be exercised when extrapolating biological responses at one level of organization to responses at another 

level. Consider a simple analogy: The sun feels good on our 

skin and it is necessary for the synthesis of vitamin D. Peel 

away the skin, expose the underlying tissue to the same 

beneficial sun, and the underlying cells die. Almost anyone 

would recognize that evaluating sunlight, based on the response of a naked cell, has nothing to do with an organism’s 

response to the same level of light. This may seem simplistic, 

but these same principles must be applied to genotoxicity 

tests. 

What question is asked by genotoxicity tests? Ernst (1994) 

reported the results of genotoxicity tests using subtidal sediments collected at various distances from a wharf constructed 

of creosote-treated wood. PAHs were extracted from the 

sediments, dried, and re-dissolved in dimethylsulfoxide 

(DMSO). Trout hepatocytes were exposed to various concentrations of the PAH preparation and genotoxicity was assayed using the nick translation assay (NTA) of Gagne and 

others (1995) and a modified version of the alkaline precipitation assay (APA) described by Olive (1988). The results 

were quantified by defining a toxicity threshold (TT) as the 

geometric mean of the lowest observed effect concentration 

(LOEC) and the no observed effects concentration (NOEC). 

This test measured the response of DNA in naked digestive 

gland cells to isolated PAHs suspended in a material, which 

is an exceptionally powerful solvent for both polar and nonpolar compounds. DMSO is often used as a reaction medium 

for bimolecular nucleophilic reactions in which the attacking 

nucleophile (arene oxide) bears a negative charge. Its use in 

these genotoxicity studies greatly facilitates transfer of PAHs 

across the plasma lemma and arene oxides into the nucleus. 

Based on this discussion, it appears that the question being 

asked is “How many DNA nicks and breaks occur when we 

eliminate, or impair, many of a cells nuclear defense mechanisms and expose DNA to PAH and their intermediate metabolic products?” This is an interesting question, and as expected, we find that the degree of DNA insult is proportional 

to the PAH exposure. In other words, this study revealed a 

quantifiable dose–response relationship. The dose is isolated 

PAHs and the response is from naked cells whose nuclear 

and cell membranes have been compromised in the presence 

of DMSO. 

Does our current understanding of bioindicators allow their 

use in assessing environmental risks? There are numerous 

weaknesses in our current understanding. Consider the following: 

Creosote and PAHs do result in disease in demersal fish at 

sufficiently high concentration levels. The review presented 

here is intended to provide insight to the mechanisms leading 

to the observed hepatic carcinomas. However, before these 

genotoxicity tests can be used to establish environmental 

criteria, we need to correlate the observed cellular responses 

with those at the population or organismal levels of organization. The response of a naked cell, with at least seven layers 

of protection stripped away, to isolated PAH, does not describe the response of whole organisms living in close association with sedimented PAH. 

Other bioindicator tests (primarily enzyme induction tests) 

suffer from the same weakness. The response of an enzyme 

system to an appropriate substrate has little to do with the 

response of the organism to that substrate. Bioindicators 

certainly have a future in environmental studies. However, 

adequate correlations between cellular or genetic responses 

and organismal or population responses to pollutant levels 

have not been made. 

Genotoxicity test 

environment 

Organismal environment 

in open aquatic systems 

PAHs are desorbed and 

extracted from sediments. 

They are made very available to the test cells. 

PAHs are bound to sediments. 

They are not readily available 

in the water column. 

No organismal epithelium 

is present. 

After desorption from sediments, PAHs must cross an 

external epithelium (skin, gills, 

gut) before entering the blood 

stream for delivery to the 

digestive gland. 

No kidney is present to 

clear PAHs. 

Kidney functions to clear some 

xenobiotics. Fish rapidly excrete most PAHs. 

Plasma lemma is compromised by DMSO. Cell membrane selectively 

restricts movement of PAH 

into the cell. This increases 

the probability of excretion and 

decreases the probability of 

metabolism. 

Lysosomal membranes 

are compromised by 

DMSO. 

Lysosomal membranes help 

contain intermediate metabolites during metabolism. 

Nuclear membrane is 

compromised by DMSO. 

Nuclear membrane provides 

another level of protection for 

DNA. 

DNA lesions are assumed 

to result in unregulated 

cell growth. 

DNA repair mechanisms 

reduce the probability of 

unregulated cell growth. 
21 

Payne and others (1988) reported a study supporting the 

hypothesis that many point sources of hydrocarbon contamination could be harmful to fish health. They found that MFO 

enzyme levels were altered at hydrocarbon levels as low as 

1.0 mg/kg. The authors noted that PAH levels in this range 

are encountered over a broad range of aquatic environments, 

many of which are not associated with pollution. They suggested that hydrocarbons often occur in sufficient concentrations to affect biological responses in fish. Consistent with 

the discussion presented here, Payne and others concluded 

that meaningful bioindicators must distinguish between 

effects per se and either chronic or acute effects. 

The development of simple, timely, and effective tests to 

evaluate the risks posed by pollutants to aquatic organisms is 

important work. However, until a better understanding of the 

correlation between effects observed in bioindicator studies 

and the response of organisms and populations of organisms 

living in open environments is achieved, bioindicators have 

little value as either regulatory or environmental health assessment tools. 

Vogelbein and others (1990) described hepatic neoplasms in 

the Mummichog (Fundulus heteroclitus) from a site with 

high levels (22 mg/kg) of PAHs in sediments. Of the Mummichogs collected at this site, 93% had gross hepatic lesions 

and 33% had hepatocellular carcinomas. Note that fish from 

a site of lower contamination (0.063 mg/kg) did not show 

signs of hepatic lesions or carcinomas. Similar cellular lesions have been described in fish from a highly urbanized 

area (Duwamish River estuary) in Puget Sound (Pierce and 

others 1977). 

Colwell (1986) examined mussels and seawater associated 

with creosoted marine pilings at the Roosevelt Roads Naval 

Station Complex in Puerto Rico. She employed Salmonella 

typhimurium in the familiar Ames test (Ames and others 

1975) for mutagenicity and found no detectable mutagenic 

activity in bacteria from either the water or mollusks associated with the creosote. She concluded that the creosote did 

not exhibit appreciable leaching into the surrounding water. 

Effects of PAHs on Aquatic Organisms 

Mesocosm studies by Stekoll and others (1980) and Widdows and others (1982,1985) reported similar community 

responses to petroleum and PAH contamination. Significant, 

long-term reductions in the abundance and diversity of invertebrate fauna were reported when ambient water levels contained as little as 130 μg/L dissolved diesel oil for prolonged 

periods. Less significant population effects were observed on 

a rocky shore community exposed to 30 ppb diesel oil for 

2 months. 

Tagatz and others (1983) examined the effect of creosotecontaminated sand on macrofaunal communities. They found 

that the lowest creosote (in sediment) concentration at either 

of their sites that affected the number of individuals or species was 844 mg/kg for mollusks and 177 mg/kg for echinoderms, annelids, and arthropods. 

The adaptation of microbial communities in the gut of Limnoria tripunctata and in sediment was well documented and 

discussed in Neff (1979). Similar adaptations were described 

by Wade and others (1989) in Gulf of Mexico hydrocarbon 

seep communities, including numerous species of annelids, 

crustaceans, bivalves, and fish. Tissue PAH concentrations 

indicate that these organisms were chronically exposed to 

high levels of PAH. The seep organisms were able to survive 

and thrive in an environment of high PAH exposure. The 

apparent ability to cope with these elevated PAH levels may 

involve specially adapted or evolved enzyme systems. 

Aquatic Toxicity Summary 

The low-molecular-weight PAHs, such as naphthalene and 

acenaphthene, produce acute toxic effects in marine animals 

because they are more soluble than the higher molecular 

weight compounds. Acute intoxication in the sensitive larval 

stages of marine invertebrates may occur at water column 

concentrations as low as 8 to 10 μg/L. However, for most 

species, the literature suggests that water column concentrations of greater that 20 μg/L are required for significant 

responses. Low-molecular-weight PAHs are more soluble 

than the high-molecular-weight compounds, and bacteria and 

other aquatic organisms more rapidly metabolize them. The 

potential for their accumulation to toxic levels is small except 

when introduced in large quantities (such as petroleum 

spills). However, laboratory (including mesocosm) studies 

have demonstrated photo-enhanced toxicity associated with 

fluoranthene and anthracene at levels as low as 3 μg/L in the 

water column. 

Because of their low biological availability, sedimented 

PAHs have a low potential to cause acute pathological responses at either the organismal or population levels in 

aquatic species. However, sediment levels of creosote exceeding 177 mg/kg have been shown to cause significant 

impacts on populations of the most sensitive taxa (Tagatz and 

others 1983). Furthermore, bacteria and eukaryotes have 

demonstrated a remarkable ability to adapt to relatively high 

levels of background PAHs. 

Chronic toxicity is more difficult to measure than acute 

toxicity. This review suggests that chronic stress can occur in 

organisms, including bivalves, at concentrations as low as 30 

to 40 μg PAH/L. Chronic stress causes reduced scope for 

growth and reduced reproductive capacity, which can have 

long-term consequences for populations of aquatic species. 

In addition to direct physiological stress (Sunda 1987), there 

is a potential for the high-molecular-weight PAHs (particularly B(a)P) to form carcinogenic, mutagenic, and teratogenic 

compounds during metabolism by crustaceans and 
22 

vertebrates. Neff (1979) summarized his section on neoplasia 

by noting that although carcinogenic PAH can produce cancer-like growths and cause teratogenesis and mutagenesis in 

some aquatic invertebrates and vertebrates, there are no 

reports of the induction of cancer by exposure of aquatic 

animals to environmentally realistic levels of carcinogenic 

PAHs in water, food, or sediments. However, more recent 

work points out increases in the number of hepatic lesions 

and carcinomas with sediment PAH burdens as low as 

10 mg/kg. 

Potential for Human Pathology 

Associated With PAHs 

Neff (1979) and Stegeman (1981) stated that consumption of 

PAH-contaminated mollusks probably constitutes a minor 

source of human dietary PAH compared with PAHs in 

smoked foods, charcoal-broiled meats, and even many vegetables. Moore and others (1989) agreed, with the caveat that 

“except possibly where animals have been exposed to very 

high concentrations of PAHs such as those occurring following an oil spill.” 

Average concentrations of 402 ppb B(a)P were reported by 

Neff (1979) in mussels from creosoted piling in harbors and 

marinas, where there are numerous sources of PAHs. Eisler 

(1987) listed human health criteria proposed by the EPA and 

others for various PAH compounds. The proposed maximum 

human consumption of benzo(a)pyrene is 1.6 μg/day. 

Goyette and Brooks (1999) compared tissue levels of PAH 

(TPAH) in mussels (Mytilus edulis) grown on and at various 

distances from new and used creosote-treated piling. Tissue 

levels of PAH peaked in samples collected 14 days following 

construction and then declined to background levels. Based 

on the highest PAH concentrations observed in these mussels, the authors determined that a person would have to eat 

408 g of mussel tissue (about 1.35 kg (3 lb) of live mussels in 

the shell) every day to exceed the U.S. EPA dietary limit for 

carcinogenic PAHs. 

Note that Mallet and others (1963) observed a B(a)P concentration of 55 μg/kg in Mytilus edulis collected from a pristine 

area on the coast of Greenland. The authors attributed this 

high level to biosynthesis of B(a)P by sediment-dwelling 

anaerobic bacteria. It was also possible that B(a)P could be 

leaching from peat deposits, which are extensive in that part 

of Greenland. Neff (1979) reported high levels of perylene 

(3.01 mg/kg) in Glastonbury peat from Avalon, England. 

This observation again points to the ubiquitous nature of 

PAHs and the multitude of natural and anthropogenic 

sources. 

The American Wood-Preservers’ Association recognizes the 

potential toxicity of creosote (AWPA 1996). The association 

provides a list of precautions to be taken in the use and handling of creosote-treated wood. This list includes precautions 

against using creosote in human or animal habitations where 

direct contact occurs. AWPA recommends that creosotetreated wood used above grade in homes be well sealed and 

that it not be used in close association with either human or 

animal feedstuffs. However, the EPA label for creosote does 

allow its incidental use, such as for piers and floats, in association with human drinking water supplies. 

Table 5 lists benchmarks for PAHs in drinking water established by the Council of European Communities (CEC), the 

Canadian Council of Ministers of the Environment (CCME), 

the British Columbia Ministry of the Environment, and the 

U.S. EPA. Of these, benzo(a)pyrene is the compound of most 

concern. 

The CEC directive relating to the quality of water 

intended for human consumption sets a maximum 

total PAH level of 200 ng/L based on the amount of 

fluoranthene, benzo(b)fluoranthene, benzo(k)fluoranthene, 

benzo(a)pyrene, benzo(ghi)perylene, and indeno(1,2,3-cd) 

pyrene. Eisler (1987) suggested that the concentration of 

these carcinogenic PAHs should be maintained at less than 

0.028 μg/L to reduce the increased cancer risk to less than 

10−5

. The concentration of the previously listed carcinogenic 

PAHs observed by Goyette and Brooks (1999) at a distance 

of 15 cm from a large creosote-treated wood structure was 

0.000051 μg/L—less than Eisler’s suggested benchmark by a 

factor of 550. The sum of all PAHs observed by Goyette and 

Brooks (1999) was 0.031 μg/L—less than the CEC benchmark by a factor of 6.5. This suggests that drinking water 

intakes could safely be placed within 15 cm of creosotetreated piling in drinking water supplies. However, that 

placement would not be recommended because it would 

represent unnecessary risk. 

Sediment Quality Benchmarks For PAHs 

Goyette and Brooks (1999) examined the effects of creosotetreated piling in Sooke Basin, British Columbia. The 

extensive physicochemical and biological database included 

infaunal community analysis and seven in situ and laboratory 

Table 5—Canadian and U.S. regulatory limits for 

drinking and ground water 

Compound Authority Remediation level 

Total carcinogenic PAH Council of 

European 

Communities 

0.200 μg/L 

Benzo(a)pyrene British Columbia 0.100 μg/L 

Benzo(a)pyrene Canadian Council 

of Ministers of the 

Environment 

0.010 μg/L 

Benzo(a)pyrene U.S. EPA 1.6 μg benzo(a) 

pyrene/day 
23 

bioassays. Physicochemical analyses included a detailed 

description of sediment and water column concentrations of 

alkylated and parental PAHs. This database allowed for an 

examination of the efficacy of existing and proposed sediment quality benchmarks in predicting adverse biological 

response. The U.S. EPA draft sediment quality criteria for 

acenaphthene (130 μg/g organic carbon), phenanthrene 

(180 μg/g organic carbon), and fluoranthene (620 μg/g organic carbon) were found to be underprotective in that they 

failed to predict observed adverse biological effects in three 

database samples. False negative responses (adverse effects 

observed but not predicted by the benchmark) were not 

observed for any of the other benchmarks. Goyette and 

Brooks (1999) found that 60 individual PAH compounds 

exceeded the threshold effects level (TEL) (Jones and others 

1997) in seven samples where no toxicity was observed. 

These false positive indications associated with the TEL 

were observed for every PAH compound except naphthalene. 

The Washington State Sediment Quality Criteria (WAC 

173–204) were most efficient in predicting adverse effects 

(12 false positive responses), and the probable effects level 

(PEL) resulted in 21 false positive responses. The mean of 

the TEL and PEL ((TEL + PEL)/2) resulted in 30 false predictions of adverse effects where none was observed. 

Swartz (1999) examined existing and proposed sediment 

quality guidelines and proposed consensus guidelines that 

appear to resolve some of the current inconsistency. He 

describes a ΣPAH toxicity threshold that is consistent with 

the effects range low (ER-L) of Long and others (1995) and a 

ΣPAH mixture LC50 that is similar to the effects range median (ER-M) described by the same authors. In the report 

herein, sediment PAH concentrations are compared with the 

ΣPAH toxicity threshold, the ΣPAH mixture LC50, and the 

mean of these two values in predicting biological risk 

at creosote-treated bridges. Table 6 summarizes these 

benchmarks. 

Site Selection 

Nine creosote-treated bridges were evaluated on the West 

Coast of the United States during the site selection process. 

Not one of these bridges was considered appropriate, primarily because the structures involved the use of both creosoteand penta-treated wood. Two bridges crossing Pipe Creek 

in Cass County, Indiana, were selected because they best 

met the site selection criteria (Fig. 7). These bridges carry 

cars, commercial vehicles, and farm equipment across a 

slow-moving, but substantial, stream. Vehicle use statistics 

were not available; however, the rural nature of the environment suggested relatively light traffic compared with urban 

roadways or interstate highways. Both bridges sit on 20 

Class A piling treated to a nominal retention of 272 kg/m3

 

(17 lb/ft3

) in the treated zone. Support beams, crossbeams, 

decking, and guardrails were all similarly treated with 

creosote oil to a retention of 128 or 160 kg/m3

 (8 or 10 lb/ft3) 

in the treated zone. 

The area is intensively used for agriculture, and Pipe Creek 

appeared to carry a significant load of topsoil (sand and silt). 

In addition, agricultural pesticides may have an effect on 

aquatic invertebrates. However, because agricultural activity 

(corn production) covers much of the landscape both upstream and downstream, it seems reasonable to suggest that 

those effects would equally influence all areas of the aquatic 

environment, including control and treatment stations. Industrial sources of PAHs were not apparent within 5 miles of 

either bridge. This qualitative assessment suggested that any 

observed differences in the invertebrate community between 

upstream controls and downstream treatment stations would 

be associated primarily with the physical presence of the 

structure or PAHs released from the creosote-treated wood, 

or both. 

Bridge 146 

Figure 8 is a photograph taken from a downstream location, 

looking upstream along the sampling transect, on November 

3, 1997, during the collection of physicochemical and biological samples at Bridge 146. The weather was overcast, 

with light wind from the north, and it was snowing. The 

water temperature was 5.5o

C, and the air temperature was 

4.5o

C at 0900.

Table 6—ΣPAH toxicity threshold, ΣPAH mixture LC50, 

and mean of these two values for 17 parental PAHsa 

PAH compound 

ΣPAH 

toxicity 

threshold 

(μg/g 

organic 

carbon) 

ΣPAH 

mixture 

LC50 

(μg/g 

organic 

carbon) 

Mean 

(μg/g 

organic 

carbon) 

Naphthalene 13 71 42.0 

Acenaphthylene 3 15 9.0 

Acenaphthene 4 23 13.5 

Fluorene 17 90 48.5 

Phenanthrene 29 155 92.0 

Anthracene 21 114 67.5 

Fluoranthene 69 371 220.0 

Pyrene 90 481 285.5 

Benz(a)anthracene 21 111 66.0 

Chrysene 31 169 100.0 

Benzo(b)fluoranthene 33 180 106.5 

Benzo(k)fluoranthene 29 155 92.0 

Benzo(a)pyrene 33 179 106.0 

Low-molecular-weight PAH 87 468 277.5 

High-molecular-weight PAH 306 1,646 976.0 

Total PAH 393 2,114 1,253.5 

a

All values, except mean, are from Swartz (1999). 
24 

Figure 8—Bridge 146 over Pipe Creek viewed from 

downstream looking upstream along the sample 

transect. Water depths varied between 10 and 31 cm. 

Water flow along the transect was <1.0 cm/s. 

Conventional physicochemical characteristics of the water 

and sediments are described in Table 7. The Indiana State 

Division of Water provided flow data for Pipe Creek. Records collected since 1960 indicated that the mean annual 

flow in Pipe Creek was 1.9 m3

/s (67.6 ft3/s). The lowest daily 

mean was 0.09 m3

/s (3.3 ft3/s) recorded February 1, 1977. 

Current speeds along the chosen transects were <1.0 cm/s, 

and water depths along the sampling transect varied between 

10 and 31 cm on November 3, 1997. 

The sampled transect was characterized by slow currents 

(<1.0 cm/s) and fine sediments dominated by sand with less 

than 10% silt and clay. Brooks (unpublished data) found that 

total organic carbon averaged 60% of the total volatile solids 

(TVS) measured in marine sediments. Extrapolating to this 

freshwater environment suggests that Pipe Creek sediments 

contained between 0.43% and 0.52% organic carbon. The 

corresponding sediment quality benchmarks for PAHs at this 

level of organic carbon are listed in Table 8. These values 

were corrected to dry sediment weight, based on the observed mean organic carbon content of 0.48%, and represent 

site-specific sediment benchmarks for this evaluation. 

Bridge 146 was constructed in 1995 and was more than 

2 years old at the time of this evaluation. The creosote model 

of Brooks (1997b) predicted continued PAH accumulation in 

sediments for about 1,000 days after construction. At the 

time of this evaluation, Bridge 146 was approximately 

900 days old, and absent any bedload movement and dispersal of PAH, the sediment concentrations should have been 

nearing their maximum concentration. However, water flow 

in Pipe Creek varies significantly with the season and it is 

likely that sediments and sediment contaminants were dispersed and carried downstream during periods of high flow. 

During this examination, local scouring was evident within 

30 cm of individual piling, and these areas were avoided in 

sampling. 

Sedimented Concentrations of PAHs 

Figure 9 compares the observed sediment concentrations of 

total PAHs with the benchmarks defined in Table 8. 

Figure 7—Location of creosote-treated bridges on Pipe Creek in Cass County, Indiana, that were subjected 

to an environmental risk assessment. 
25 

Figure 9—Sediment concentrations of total PAH 

measured upstream (controls) and downstream from 

Bridge 146 crossing Pipe Creek in Cass County, Indiana. 

Sediment PAH concentrations are in micrograms per 

gram (dry sediment weight). The threshold effects level 

(TEL), mean effects level ((TEL + PEL)/2), and probable 

effects level (PEL) are computed at the observed level of 

total organic carbon (0.48%). 

Table 7—Physicochemical characteristics observed at Bridge 146 on Pipe Creek, Cass County, Indiana, 

November 3, 1997a 

Station 

Depth 

(cm) 

RPD 

(cm)b

Total PAH 

(μg/g) 

TVSc

 

(%) 

Sand 

 (%) 

Silt and clay 

(%) 

−22.8 m (−76 ft) upstream control 10.0 >3.0 0.2288 NDd

 1.98 + 1.22 81.03 + 10.22 14.37 + 3.21 

0.0 under bridge 24.0 >3.0 1.2801 + 0.8612 0.86 + 0.22 83.08 + 26.19 5.84 + 0.63 

0.45 m (1.5 ft) downstream 21.0 >3.0 2.1396 0.77 72.91 6.31 

0.9 m (3.0 ft) downstream 15.0 >3.0 2.8426 0.72 90.88 7.67 

1.8 m (6.0 ft) downstream 17.0 >3.0 5.5376 1.01 91.39 4.59 

3 m (10.0 ft) downstream 16.0 >3.0 1.7476 0.80 89.59 7.15 

6 m (20.0 ft) downstream 10.5 >3.0 0.5402 0.83 48.03 7.00 

22.8 m (33.0 ft) downstream 20.0 >3.0 0.2288 ND 0.72 60.70 3.36 

a

All values are the mean of two replicates except the −22.8 m (−76 ft) upstream control and samples taken under the 

 bridge, which are the mean of three replicates. Where appropriate, values include 95% confidence limits on the mean.

b

RPD is the reduction–oxidation potential discontinuity. It is an indication of the depth in sediments at which sufficient 

 oxygen is present to support aerobic metabolism. 

c

TVS is total volatile solids measured as a percentage of sediment dry weight. 

d

PAHs were not detected. The value is the sum of the detection limits. 

Table 8—Consensus sediment quality benchmarks 

specific to total organic carbon observed at Bridge 

146 over Pipe Creek in Cass County, Indianaa 

PAH compound 

ΣPAH 

toxicity 

threshold 

ΣPAH 

Mixture 

LC50 Mean 

Naphthalene 0.06 0.34 0.20 

Acenaphthylene 0.01 0.07 0.04 

Acenaphthene 0.02 0.11 0.06 

Fluorene 0.08 0.43 0.23 

Phenanthrene 0.14 0.74 0.44 

Anthracene 0.10 0.55 0.32 

Fluoranthene 0.33 1.78 1.06 

Pyrene 0.43 2.31 1.37 

Benz(a)anthracene 0.10 0.53 0.32 

Chrysene 0.15 0.81 0.48 

Benzo(b)fluoranthene 0.16 0.86 0.51 

Benzo(k)fluoranthene 0.14 0.74 0.44 

Benzo(a)pyrene 0.16 0.86 0.51 

Low molecular weight PAH 0.42 2.25 1.33 

High molecular weight PAH 1.47 7.90 4.68 

Total PAH 1.89 10.15 6.02 

a

Swartz (1999). All values are μg PAH/g dry sediment 

 weight. 

0 

2 

4 

6 

8 

10 

12 

Current direction 

-22.8 0 0.45 0.9 1.8 3.0 6.0 9.9 

Distance from Bridge146 (m)

PAH concentration (

μg/g) 

Total PAH 

TPAH TEL 

TPAH MEL 

TPAH PEL 
26 

Goyette and Brooks (1999) provided a detailed history of 

individual PAH compounds present in new creosote oil, 

treated piling, the water column, and sediments at Sooke 

Basin as a function of time following installation. They found 

that much of the low-molecular-weight compounds were lost 

during the treating process or were dissolved in the water 

column and diluted. The intermediate weight compounds, 

phenanthrene through chrysene, presented the greatest risk to 

aquatic organisms, and the high-molecular-weight compounds were found only in very low concentrations. Table 8 

summarizes the ΣPAH toxicity threshold (TEL), ΣPAH 

mixture LC50 (PEL), and the mean of these two values ((TEL 

+ PEL)/2) for 17 parental PAHs adjusted to the mean sediment organic carbon (0.48%) determined November 3, 1997. 

These were site-specific benchmarks for predicting biological response at Bridge 146. 

Observed PAH concentrations in surficial sediment samples 

(upper 2.0 cm of the sediment column) are described in 

Table 9 for Bridge 146. Concentrations exceeding the TEL 

(ΣPAH toxicity threshold) were observed at distances between 0.45 and 1.8 m (1.5 and 6.0 ft) from the downstream 

perimeter of the bridge structure. Nearly all the low and 

intermediate molecular weight compounds exceeded their 

individual TELs at these same distances. Not one of the 

concentrations of high-molecular-weight compounds was 

appreciably elevated, and those for which Swartz published 

data were not elevated above their TEL. All three replicate 

sediment samples collected under the bridge (0.0 m) contained significantly elevated (α = 0.05) levels of naphthalene 

and acenaphthene compared with the upstream control. 

These levels exceeded both the (TEL + PEL)/2 level and the 

PEL. Acenaphthene and phenanthrene concentrations exceeded the PEL at the station located 1.8 m (6.0 ft) downstream from the bridge. If the PAHs were bioavailable, these 

levels could result in significant acute biological effects in 

sensitive species. The low-molecular-weight PAHs (naphthalene and acenaphthene) have higher water solubilities than 

the higher molecular weight compounds, are therefore more 

biologically available, and are generally not associated with 

chronic effects. Their catabolism does not result in the production of arene oxides and other cancer-promoting intermediate metabolites. These results suggest that acute effects 

may be present in aquatic organisms under the bridge and to 

distances of 1.8 m (6.0 ft) downstream. 

Table 9—Concentrations of sedimented PAHs observed in the vicinity of Bridge 146, Cass County, Indianaa 

Concentrations (μg/g dry sediment weight) for various sampling stations 

Compound 

Reporting 

limit 

(μg/g) 

−22.8 m 

(−76 ft) 

0−0.3 m 

(0.0−1) 

+0.45 m 

(+1.5) 

+0.9 m 

(+3.0) 

+1.8 m 

(+6.0) 

+3 m 

(+10) 

+6 m 

(+20) 

+9.9 m 

(+33 ft) 

Naphthalene 0.020 0.010 XX 0.617 0.042 0.061 0.180 0.066 0.010 0.010 

Acenaphthylene 0.020 0.010 0.014 0.010 0.010 0.010 0.010 0.010 0.010 

Acenaphthene 0.020 0.010 XX 0.163 0.051 0.065 XX 0.310 0.096 0.026 0.010 

Fluorene 0.010 0.010 0.060 0.066 0.083 0.360 0.099 0.022 0.010 

Phenanthrene 0.020 0.010 0.120 0.330 0.330 XX 1.600 0.460 0.092 0.010 

Anthracene 0.020 0.010 0.023 0.079 0.067 0.010 0.059 0.024 0.010 

Fluoranthene 0.020 0.010 0.095 0.560 0.900 1.200 0.390 0.110 0.010 

Pyrene 0.020 0.010 0.077 0.500 0.770 0.950 0.340 0.097 0.010 

Benzo(a)anthracene 0.003 0.001 0.032 0.170 0.170 0.270 0.078 0.063 0.001 

Chrysene 0.020 0.010 0.010 0.120 0.170 0.190 0.010 0.010 0.010 

Benzo(b)fluoranthene 0.004 0.002 0.002 0.067 0.077 0.083 0.040 0.002 0.002 

Benzo(k)fluoranthene 0.003 0.002 0.002 0.033 0.043 0.054 0.015 0.002 0.002 

Benzo(a)pyrene 0.005 0.002 0.002 0.063 0.052 0.086 0.030 0.002 0.002 

Benzo(ghi)perylene 0.010 0.010 0.010 0.024 0.010 0.010 0.010 0.010 0.010 

Ideno(1,2,3-cd)pyrene 0.009 0.004 0.004 0.004 0.004 0.004 0.004 0.004 0.004 

Dibenzo(a,h)anthracene 0.006 0.003 0.003 0.003 0.003 0.003 0.003 0.003 0.003 

Total PAH 0.209 0.114 1.236 2.122 2.815 5.320 1.710 0.487 0.114 

a

Reported concentrations are from the top 2 cm of the sediment column. Sediment PAH concentrations exceeding the 

 threshold effects level, defined by Swartz (1999) as the ΣPAH toxicity threshold, are in bold type. Sediment PAH 

 concentrations exceeded the probable effects level, defined by Swartz (1999) as the ΣPAH mixture LC50, are preceded 

 by XX. Cells containing sediment PAH concentrations exceeding the (TEL + PEL)/2 benchmark used in this assessment 

 have a gray background. 
27 

However, Long and others (1998) examined the efficiency of 

TEL and PEL in predicting toxicity. They found that values 

between TEL and PEL did not predict increased toxicity 

(over controls) until the number of levels exceeding TEL was 

more than 20. A similar analysis using PEL suggested that 

53.3% of the samples were judged not toxic when one to five 

PEL were exceeded. The nontoxic samples decreased to 37% 

when 6 to 10 PELs were exceeded, and the benchmark’s 

efficiency further increased to 10% nontoxic responses when 

11 to 20 PELs were exceeded in a mixture. In this case, two 

PELs were exceeded under the bridge and at +1.8 m (+6.0 ft) 

downstream. For this number of exceedances, Long and 

others (1998) found that 52% of the samples were nontoxic, 

12% were marginally toxic, and 36% were highly toxic. 

To put these values in better perspective, Long and others 

(1998) reported that 65% of the sediment samples in which 

TELs were not exceeded tested nontoxic. However, 26% of 

these relatively clean sediments were found to be marginally 

toxic and 9% were highly toxic. Goyette and Brooks (1999) 

found that the Sooke Basin Reference Station was consistently toxic in a suite of bioassay tests compared with reference sediments from Esquimalt Lagoon. Similarly, Deniseger 

and Erickson (1998) found that 8 of 12 reference sites evaluated using Eohaustorius washingtonianus in the Broughton 

Archipelago were toxic when the results were compared with 

sediments from which the test animals were collected. There 

are numerous site-specific physicochemical and biological 

factors that determine the ultimate toxicity of contaminants. 

The benchmarks (TEL, PEL, (TEL + PEL)/2) against which 

Pipe Creek sediment concentrations of PAHs were compared 

are only as good as the screening tools. Based on the results 

of this analysis, there is a reasonable likelihood that sedimented PAHs under Bridge 146 and at a distance of 1.8 m 

(6.0 ft) downstream could be associated with moderate adverse biological effects. The information displayed in Table 

9 is provided in graphical format in Figure 9. It is apparent 

that PAHs were either being transported downstream prior to 

being deposited in sediments or that the PAHs were deposited under the bridge, then moved downstream with the 

sediment bedload. In the latter case, the PAH profile presented in Figure 9 represents a time history of PAH lost (after 

degradation) from the bridge. This study was not designed to 

assess PAH transport from creosote-treated structures, and 

this observation is simply a hypothesis worthy of additional 

testing. The observed spatial profile is more likely a result of 

both factors. 

Biological Response—Infauna 

A total of 8,187 invertebrates were identified in 18 sediment 

samples collected in the vicinity of Bridge 146. Each sample 

represented an area of 0.0309 m2

. These samples contained 

51 taxa. Taxonomic codes are defined in Appendixes B and 

C. Dominant taxa were defined as those representing 1% of 

the total infaunal abundance (count = 82). The invertebrate 

community was dominated by annelids (4,579 total) and 

chironomids in the genus Chironomus (3,107). Subdominant 

taxa were defined as those representing 1% of the abundance 

of all taxa except the two dominants (count = 5.0). In addition, to be included in the subdominant database, a taxon 

must have been found in at least 10% of the samples (2.0 

samples). Eighteen taxa met these criteria. Dominant and 

subdominant taxa represented 99.2% of the total taxa and are 

identified in Table 10. 

This was a robust community dominated by annelids and 

chironomids. Few mayflies (Order Ephemeroptera) or 

cadisflies (Order Trichoptera) and no stoneflies (Order 

Table 10—Dominant and subdominant invertebrate taxa 

observed in sediment samples collected in the vicinity of 

Bridge 146 crossing Pipe Creek in Cass County, Indianaa 

Taxa Code 

Found in 

number of 

samples 

Total 

abundance 

Dominant Taxa 

Annelids ANNE 18 4,579 

Chironomus sp. AICHSS 18 3,107 

Sub-dominant Taxa 

Gastropods 

Pleurocera sp. MGPLS 10 36 

Ferrisia sp. MGFRS 7 9 

Physella sp. MGPHL

S 

 9 54 

Ostracods ACOST 2 11 

Order Ephemeroptera 

Caenis sp. AICAE 14 46 

Ephemera sp. AIEPHA 8 24 

Order Hemiptera 

Family Corixidae AICOR 6 12 

Order Coleoptera 

Berosus sp. AIBERS 5 7 

Dubiraphia sp. AIDUBS 13 49 

Order Trichoptera 

Polycentropus sp. AIPOSS 4 5 

Order Diptera 

cf. Probezzia sp. AIPBEZ 7 30 

Family Chironomidae AICHR 9 40 

Procladius sp. AIPRS 5 7 

Dicrotendipes sp. AIDTS 16 79 

Eukiefferiella sp. AEUKS 7 13 

cf. Chrysops sp. ACHS 6 14 

Total abundance of dominants and subdominant 

taxa 

8,122 

a

Eighteen samples, each covering 0.0309 m2, were collected. 
28 

Plecoptera) were observed in the fine-grained sediments 

characteristic of this part of Pipe Creek. The composition of 

the infaunal community is undoubtedly influenced by soil 

loss from adjacent agricultural areas and subsequent movement of this bedload of sand, silt, and clay down the watershed. In Figure 10, total abundance, subdominant abundance, 

and the number of taxa observed in 0.0309-m2

 samples are 

compared with sediment PAH concentrations. More taxa 

were observed under and downstream from the bridge than at 

either the upstream control or the downstream-most station, 

where PAHs were not detected in sediments above the analytical detection limit. The abundance of all taxa was slightly 

reduced under the bridge and a distance of 0.45 m (1.5 ft) 

downstream. Total abundance was highest at the 1.8-m 

(6.0-ft) station where TPAH concentrations peaked at 

5.3 μg/g. Subdominant taxon abundance decreased at stations 

located under the bridge and 1.8 m downstream. However, 

subdominant taxa were more abundant at all downstream 

stations than at the upstream control. 

Figure 11 does not suggest significant differences in either 

Shannon’s Index or Pielou’s Index as a function of either 

sediment PAH concentration or downstream distance from 

the bridge. The simplest interpretation of Pielou’s Index is 

that it represents the proportion of the maximum value of 

Shannon’s Index observed in a sample. All calculated values 

of Shannon’s Index were low (about 1.0), indicating the 

stressful nature of the environment. Similarly, Pielou’s Index 

was less than 0.5 in all samples. 

PAH data (Table 9) suggest two areas where adverse biological effects might be observed. The area under the bridge 

contained high levels of the low-molecular-weight compounds naphthalene and acenaphthylene. In addition, the 

station located 1.8 m downstream from the downstream 

perimeter of the bridge contained levels of acenaphthene and 

phenanthrene exceeding the PEL. 

The significance of biological effects was explored using ttests on log(n + 1) transformed count data and log transformed variables with continuous distributions. These transformations were needed to meet the underlying assumptions 

for parametric testing. Significant (α = 0.05) differences in 

biological response variables (number of taxa, abundance, 

subdominant abundance, and Shannon’s Index) were not 

observed between the upstream control and either the 0.0- or 

1.8-m downstream stations where adverse effects might be 

associated with elevated concentrations of creosote-derived 

PAH. The null hypothesis that response variables at the 

upstream control compared with samples from under the 

bridge or at a distance of 1.8-m downstream were significantly different was not rejected with the probability of being 

equal, ranging from p = 0.58 to 0.76. 

The biological response to sedimented PAHs was further 

explored using principal components analysis (PCA). The 

factors were rotated using varimax normalization, and the 

results are presented in Figure 12. Factor loadings on factor 1 

were all positive except total PAH, which was close to zero 

but negative (−0.033886). Annelids (LANNE), the mayfly, 

Ephemera sp. (AIEPHA), and the biological metrics, taxa 

(LTAXA), abundance (LABUNDAN), and subdominant 

abundance (LDABUND) were significantly loaded on 

factor 1. Total PAH (LTPAH) was significantly loaded on 

the positive end of factor 2; many individual taxa were 

negatively loaded on this axis, but only the dipteran, 

Probezzia sp., was a significant factor. Subtle ecological 

Figure 10—Abundance of all taxa, subdominant taxa, 

and sediment levels of total PAH (μg/g dry sediment 

weight) observed in the vicinity of Bridge 146 crossing 

Pipe Creek in Cass County, Indiana, November 3, 1997. 

Figure 11 —Comparison of Shannon’s and Pielou’s 

Indexes with total PAH (TPAH) observed in sediments at 

Bridge 146 crossing Pipe Creek in Cass County, Indiana, 

November 3, 1997. 

0 

200 

400 

600

800

1000

1200

Distance from Bridge 146 (m)

Total abundance (number/0.0309 m2

)

0 

5 

10 

15 

20 

25 

30 

35 

40 

TPAH (

μg/g) and subdominant 

abundance and number of taxa 

Total abundance 

TPAH 

Subdominant 

abundance

Number taxa 

-22.8 0 0.45 0.9 1.8 3.0 6.0 9.9 0

1

2

3

4

5

6

Distance from Bridge 146 (m)

TPAH (

μg/g) 

Shannon's Index and Pielou's Index

TPAH

Shannon's Index

Pielou's Index

-22.8 0 0.45 0.9 1.8 3.0 6.0 9.9 

Current direction 
29 

Figure 12—Principal components analysis of log(n + 1) transformed biological 

response variables and log transformed PAH concentrations in samples 

collected from Pipe Creek in the vicinity of Bridge 146 in Cass County, Indiana, 

November 3, 1997. 

relationships and possible effects associated with sedimented 

PAHs are shown in Figure 12. These include the following: 

• The number of taxa and both total abundance (LABUNDAN) and subdominant species abundance (LDABUND) 

group together with annelids (LANNE) and mayflies in the 

genus Ephemera (LAIEPHA). This suggests that more 

abundant and diverse communities were found in association with annelids and mayflies. Shannon’s Index describes 

the evenness of the distribution of animals within species 

in a community. Note that Shannon’s Index (LSHANNON) is located well to the left of the other biological 

endpoints, suggesting that a few species tend to dominate 

Pipe Creek infauna and that increases in the more abundant 

taxa resulted in decreases in Shannon’s Index. 

• All the PAH group in the upper left corner of the chart 

(Total PAH = LTPAH, naphthalene = LNAPH, acenaphthene = LACENAPH, and phenanthrene = LPHENAN). The concentration of total PAH and/or naphthalene 

may have been more stressful than either acenaphthene or 

phenanthrene. 

• The presence of significant quantities of naphthalene is 

unexpected with creosote contamination. Goyette and 

Brooks (1999) found that most of the naphthalene was lost 

during the treating process from piling produced using Best 

Management Practices (WWPI 1996). This process requires the use of “clean creosote” containing less than 

0.5% xylene-insoluble material (<1.5% when treating ponderosa or southern yellow pine). In addition, these procedures require heating of the treated wood for 1 h in an expansion bath under 74 kPa (22 inHg) of vacuum for 2 h 

following impregnation. This is followed by a final steaming at <116o

C for 2 h (timber) or 3 h (piling). Following 

these steps, the treated wood is placed under a minimum 

74-kPa (22-inHg) vacuum for 4 h. The Best Management 

Practices process releases excess preservative from the 

treated wood and results in a product with a minimum 

amount of surface residue. Naphthalene has a relatively 

low vapor pressure and is more soluble in water than the 

other PAHs. It is likely that the Best Management Practices process caused the significant loss of naphthalene observed by Goyette and Brooks (1999). Because of its age, 

it is unlikely that the wood used to construct these bridges 

was treated using Best Management Practices procedures. 

 

Factor 1 

Factor 2 

LANNE 

 

LMGPLS 

 

LMGPHLS 

LAICAE 

LAIEPHA 

 

LAIPBEZ 

 

LAICHR 

 

LAICHSS 

 

LAIDTS 

 LTAXA 

 

LABUNDAN 

 

LDABUND 

 

LSHANNON 

 

LTPAH 

 

LNAPH 

 LPHENA N 

LACENAPH 

- 1.0 

 

- 0.6 

 

- 0.2 

 

0.2 

 

0.6 

 

1.0 

 

- 0.1 

 0.1  0.3  0.5  0.7  0.9  1.1  

LAIDUBS 
30 

• The taxa present in Pipe Creek were not particularly sensitive to PAHs in the vicinity of Bridge 146. Correlation 

analysis on the log-transformed data was used to investigate the relationship between total PAHs, naphthalene, 

acenaphthene, phenanthrene, and biological endpoints. All 

significant correlations with individual PAHs were positive, indicating increased abundance in the presence of 

higher sediment concentrations of naphthalene, acenaphthene, and phenanthrene. Significant negative correlations were found between the gastropod Physella sp. 

(MGPHLS, r = −0.59) and the dipteran Bezzia sp. (LAIPBEZ, r = −0.71)). A significant positive correlation was 

found between LTPAH and the mayfly Caenis sp. (AICAE, r = 0.68). Chironomids in the Family Chironomidae 

(AICHR, r = 0.27) and the dominant genus Chironomus 

sp. (AICHSS, r = 0.17) were also tolerant of elevated 

PAHs, but the correlations were not significant. The correlation analysis is summarized in Table 11. 

Sediment Bioassay Response 

Ten-day sediment toxicity tests were completed using the 

amphipod Hyalella azteca in accordance with ASTM E1706 

(ASTM 1997) as modified by MEC Analytical Systems 

Protocol P034.0. Five tests were rejected during the initial 

testing series as a result of low control survival. MEC attributed this to stressed amphipods used for that series. 

New sediments were collected at B146 and B148 in August 

1998, and the tests were repeated. Two levels of control were 

provided in these bioassays. Standard MEC laboratory test 

sediments were used as the first level of control. A second 

level of control was provided by evaluating the results of 

treatment stations located at +0.45, +1.8, and +9.9 m (+1.5, 

+6.0, and +33 ft) downstream with sediments from the local 

control station located 22.8 m (76 ft) upstream at B146. 

Survival was excellent in all tests, and significant differences 

were not observed in t-tests on log-transformed proportions 

between any of the test stations. There were no significant 

differences in survival, and the null hypothesis that survival 

was equal between any pair of stations was not rejected. 

Risk Assessment Summary 

Moderately elevated PAH levels were observed in sediments 

under and immediately downstream from Bridge 146. Sediment levels of naphthalene and acenaphthylene exceeded the 

PEL given by Swartz (1999) in sediments collected under the 

bridge. Sediment levels of PAH peaked 1.8 m (6.0 ft) downstream from the downstream perimeter of the bridge. The 

sum of all priority pollutant PAHs exceeded the TEL of 

Swartz (1999) but did not exceed the PEL or the intermediate 

value ((TEL + PEL)/2) invoked as a benchmark for this 

evaluation. Sediment concentrations of acenaphthene and 

phenanthrene exceeded the PEL at station 1.8 m (6.0 ft). The 

weight of evidence suggested that minor adverse effects 

could be observed under these conditions. 

Pipe Creek flows through an agricultural landscape devoted 

primarily to the production of corn. The stream carries a 

significant bedload of sand, silt, and clay. The resulting 

invertebrate community is dominated by annelids and chironomids. This community did not include significant numbers of larvae in the more sensitive Orders Ephemeroptera, 

Plecoptera, or Trichoptera. Pipe Creek invertebrates constituted a robust community living in a stressful environment. 

Significant differences in the abundance of organisms, the 

number of taxa or their diversity, as measured by Shannon’s 

Index, were not observed. However, the number of taxa and 

the abundance of subdominant taxa were observed to decline 

under the bridge and at the +1.8-m (+6.0-ft) station where 

exceedances of the PEL were observed. These subtle effects 

were further examined and their reality substantiated using 

principle components analysis. However, it must be emphasized that the observed differences were small and not statistically significant. The observations could have been simply 

the result of random sampling. 

Ten-day sediment bioassays using the amphipod Hyalella 

azteca did not reveal significant survival differences in sediments from stations located downstream from the bridge 

when compared with those located upstream or with 

laboratory control sediments. 

Table 11—Summary of Pearson correlation coefficients 

between individual taxa (Column 1) and total PAH 

(LTPAH), naphthalene (LNAPH), phenanthrene 

(LPHENAN), and acenaphthene (LACENAPH)a 

 LTPAH LNAPH LPHENAN LACENAPH 

LANNE −0.26 0.27 0.46 0.50 

LMGPLS 0.40 0.21 0.73 0.52 

LMGPHLS −0.59 −0.14 −0.19 −0.11 

LAICAE 0.68 0.52 0.61 0.49 

LAIEPHA 0.10 0.29 0.52 0.40 

LAICOR 0.41 0.39 0.24 0.17 

LAIDUBS −0.56 −0.12 −0.49 −0.26 

LAIPBEZ −0.71 −0.53 −0.31 −0.49 

LAICHR 0.27 −0.09 0.65 0.29 

LAICHSS 0.17 0.18 0.28 0.32 

LAIDTS −0.01 0.27 0.25 0.23 

LAEUKS 0.32 0.68 0.39 0.61 

LACHS −0.44 −0.04 −0.32 −0.07 

LTAXA 0.25 0.12 0.54 0.30 

LABUNDAN −0.19 0.23 0.48 0.49 

LDABUND −0.38 0.06 0.25 0.14 

LSHANNON 0.12 0.19 0.10 0.03 

a

Taxa codes are listed in Table 10. 
31 

Bridge 148 

The photograph in Figure 13 was taken November 4, 1997, 

from a downstream location, looking upstream. The photograph in Figure 14 was taken from a position under 

Bridge 148 looking downstream along the sampling transect. 

The weather was overcast, with light wind from the northwest, and it was snowing. 

The water temperature was 5.5o

C, and the air temperature 

was 5.5o

C at 1200 hours. Conventional physicochemical 

characteristics measured in water and sediments are described in Table 12. The Indiana State Division of Water 

provided flow data for Pipe Creek. Records collected since 

1960 indicate that the mean annual flow in Pipe Creek is 

1.91 m3

/s (67.6 ft3/s). The lowest daily mean was 0.09 m3/s 

(3.3 ft3

/s) recorded on February 1, 1977. Current speeds 

along the sampling transect were too slow to be measured 

with either a drogue or a current meter equipped with a 

magnetic head. Current speeds in the main channel varied 

between 0.16 and 0.27 m3

/s (5.83 and 9.67 ft3/s) in water 

depths of 20 to 50 cm. 

Sediment Concentrations of PAHs 

Mean total volatile solid (TVS) at stations 0.0 to 20.0 was 

1.48 + 0.66. Using the conversion factor of 0.6 (Brooks, 

unpublished data) suggests a total organic content of 0.89%. 

Table 13 defines sediment PAH benchmarks for Bridge 148 

derived from Swartz (1999) at 0.89% total organic carbon. 

Figure 15 compares sediment concentrations of PAHs observed at Bridge 148 with these benchmarks; Table 14 summarizes PAH data. Not one of the benchmarks was exceeded, 

Figure 13—Bridge 148 over Pipe Creek viewed from 

downstream. 

Figure 14—Bridge 148 looking downstream along the 

chosen sample transect 

Table 12—Physicochemical characteristics observed at Bridge 148 on Pipe Creek, Cass County, Indiana, 

on November 4, 1997a 

Station 

Depth 

(cm) 

RPDb

(cm) 

Total PAH 

(μg/g) 

TVSc 

(%) 

Sand 

(%) 

Silt and clay 

(%) 

−9 m (–30 ft) upstream control 17 1.5 0.1144 (NDd

) 0.91 + 0.21 72.17 + 16.67 8.01 + 3.14 

0.0 under bridge 29 >3.0 0.9813 + 3.1781 1.15 + 1.05 89.46 + 7.52 8.48 + 5.52 

0.45 m (1.5 ft) downstream 19 >3.0 1.9573 1.14 88.37 7.83 

0.9 m (3.0 ft) downstream 14 >3.0 0.6414 1.10 89.04 9.83 

1.8 m (6.0 ft) downstream 12 >2.0 2.2563 2.22 85.71 13.09 

3 m (10 ft) downstream 23 >3.0 0.5654 1.71 88.74 9.80 

6 m (20 ft) downstream 24 >3.0 0.3194 1.56 85.68 9.21 

9.9 m (33 ft) downstream 14 >3.0 0.5160 5.91 27.50 70.40 

a

All values are the mean of two replicates except the −9 m (–30 ft) upstream control and samples taken under the 

 bridge, which are the mean of three replicates. Where appropriate, values include 95% confidence limits on the 

 mean. Total PAH values include half the detection limit when individual PAHs were not observed. 

b

RPD is the reduction–oxidation potential discontinuity. It is an indication of the depth in sediments at which 

 sufficient oxygen is present to support aerobic metabolism. 

c

TVS is total volatile solids measured as a percentage of dry sediment weight. 

d

PAHs were not detected. The value is the sum of these detection limits. 
32 

Figure 15—Sediment concentrations of total PAH 

measured upstream (controls) and downstream from 

Bridge 148 crossing Pipe Creek in Cass County, Indiana. 

Sediment PAH concentrations are in micrograms per 

gram (dry sediment weight). The threshold effects level 

(TEL), mean effects level ((TEL + PEL)/2), and probable 

effects level (PEL) are computed at the observed level of 

total organic carbon (0.89%). 

Table 13—Consensus sediment quality benchmarks 

specific to the total organic carbon observed at Bridge 

148 over Pipe Creek in Cass County, Indianaa 

PAH compound 

ΣPAH 

toxicity 

threshold 

ΣPAH 

mixture 

LC50 Mean 

Naphthalene 0.12 0.63 0.37 

Acenaphthylene 0.03 0.13 0.08 

Acenaphthene 0.04 0.20 0.12 

Fluorene 0.15 0.80 0.48 

Phenanthrene 0.26 1.38 0.82 

Anthracene 0.19 1.01 0.60 

Fluoranthene 0.61 3.30 1.96 

Pyrene 0.80 4.28 2.54 

Benz(a)anthracene 0.19 0.99 0.59 

Chrysene 0.28 1.50 0.89 

Benzo(b)fluoranthene 0.29 1.60 0.95 

Benzo(k)fluoranthene 0.26 1.38 0.82 

Benzo(a)pyrene 0.29 1.59 0.94 

Low-molecular-weight PAH 0.77 4.17 2.47 

High-molecular-weight PAH 2.72 14.65 8.69 

Total PAH 3.50 18.81 11.16 

a

Swartz (1999). 

Table 14—Concentrations of sedimented PAHs observed in the vicinity of Bridge 148, Cass County, Indianaa

Concentrations (μg/g dry sediment weight) for various sampling stations 

Compound 

Reporting 

limit 

(μg/g) 

–9.9 m 

(–33 ft) 

0–0.3 m 

(0–1 ft) 

+0.45 m 

(+1.5 ft) 

+0.9 m 

(+3.0 ft) 

+1.8 m 

(+6.0 ft) 

+3 m 

(+10 ft) 

+6 m 

(+20 ft) 

9.9 m 

(+33 ft) 

Naphthalene 0.020 0.010 0.0260 0.0100 0.0520 0.0610 0.0100 0.0220 0.0230 

Acenaphthylene 0.020 0.010 0.0100 0.0100 0.0100 0.0100 0.0100 0.0100 0.0230 

Acenaphthene 0.020 0.010 0.0100 0.0100 0.0380 0.0930 0.0210 0.0100 0.0230 

Fluorene 0.010 0.010 0.0150 0.0690 0.0100 0.1300 0.0260 0.0100 0.0630 

Phenanthrene 0.020 0.010 0.2367 0.6100 0.1300 0.5900 0.1000 0.0590 0.0040 

Anthracene 0.020 0.010 0.0283 0.0700 0.0220 0.0890 0.0240 0.0100 0.0040 

Fluoranthene 0.020 0.010 0.3450 0.4900 0.1200 0.4800 0.1300 0.0750 0.0055 

Pyrene 0.020 0.010 0.2690 0.4200 0.0970 0.4600 0.1400 0.0890 0.0230 

Benzo(a)anthracene 0.003 0.001 0.0082 0.1300 0.0490 0.1400 0.0540 0.0013 0.0230 

Chrysene 0.020 0.010 0.0100 0.0100 0.0100 0.0640 0.0100 0.0100 0.0070 

Benzo(b)fluoranthene 0.004 0.002 0.0018 0.0420 0.0018 0.0460 0.0018 0.0018 0.1300 

Benzo(k)fluoranthene 0.003 0.002 0.0017 0.0250 0.0820 0.0260 0.0190 0.0017 0.0230 

Benzo(a)pyrene 0.005 0.002 0.0023 0.0440 0.0023 0.0500 0.0023 0.0023 0.0100 

Benzo(ghi)perylene 0.010 0.010 0.0100 0.0100 0.0100 0.0100 0.0100 0.0100 0.0230 

Ideno(1,2,3-cd)pyrene 0.009 0.004 0.0043 0.0043 0.0043 0.0043 0.0043 0.0043 0.0230 

Dibenzo(a,h)anthracene 0.006 0.003 0.0030 0.0030 0.0030 0.0030 0.0030 0.0030 0.1400 

Total PAH 0.209 0.114 0.9813 1.9573 0.6414 2.2563 0.5654 0.3194 

a

Reported concentrations are from the top 2 cm of the sediment column. Sediment PAH concentrations exceeding the 

 threshold effects Level, defined by Swartz (1999) as the ΣPAH toxicity threshold, are in bold type. Cells containing 

 sediment PAH concentrations exceeding the (TEL + PEL)/2 benchmark used in this assessment have a gray background. 

0 

 

 

 

2 

 

 

 

4 

 

 

 

6 

 

 

 

8 

 

 

 

10 

 

 

 

12 

 

 

 

14 

 

 

 

16 

 

 

 

18 

 

 

 

20 

 

 

 

- 9.9 

 

 

 0 

 

 

 0.45 

 

 

 

0.9 

 

 

 1.8 

 

 

 3.0 

 

 

 

6.0 

 

 9.9 

Distance from Bridge 148 (m) 

Total PAH/g dry sediment (

μg)

Total PAH 

 

 

 TPAH TEL 

 

 

 TPAH MEL 

 

 

 TPAH PEL 

 

 

 
33 

including the threshold effects level, and no biological effects 

should be associated with sediment levels of PAHs observed 

at this bridge. As with Bridge 146, highest sediment concentrations of PAHs were observed in the sample collected 

1.8- m (6.0-ft) downstream from the downstream drip-line of 

the bridge. This suggests that PAHs from the bridge are 

deposited in this area as a function of water depth, PAH 

particle size, and current speed. 

Goyette and Brooks (1999) observed a shift toward increasing proportions of high-molecular-weight PAHs in sediments 

at the Sooke Basin Creosote Evaluation Study site. Unlike 

Pipe Creek, currents in Sooke Basin were very slow and 

there was no evidence suggesting that PAHs, when sedimented, were redistributed by storms or tides. The low and

intermediate weight compounds (through pyrene) appeared 

to be fairly rapidly degraded, likely by bacteria, in Sooke 

Basin sediments. The higher molecular weight compounds 

are more refractory to microbial degradation, and their proportion likely increased as a result of the hypothesized differential catabolism. 

Bridge 148, constructed in 1980, was 17 years old during this 

evaluation. In contrast, Bridge 146, constructed in 1995, was 

only 2 years old. If PAHs lost from these bridges remained 

undisturbed in Pipe Creek sediments, then a similar shift to a 

higher proportion of high-molecular-weight compounds 

would have been expected. In fact, as depicted in Figure 16, 

the suite of sedimented PAH compounds are remarkably 

similar at the two bridges. Characteristic of creosote, phenanthrene, fluoranthene, and pyrene were dominant compounds 

at both bridges. The high-molecular-weight compounds 

remained at relatively low concentrations. This suggests that 

sedimented PAHs observed in this study were recently deposited (within the past year) and that historic deposits, in 

which low and intermediate compounds were degraded 

leaving a dominance of high-molecular-weight compounds, 

were likely redistributed and diluted as sediments moved 

downstream during winter and spring high water flows. If 

this hypothesis is correct, then reduced PAH concentrations 

would be expected each spring. PAHs would be released 

from the bridges during hot summer weather and they would 

accumulate in downstream sediments, reaching a maximum 

during the fall and early winter before high flows reduce their 

concentration levels. This study was conducted in the late fall 

while stream flows were still low. If this hypothesis is correct, then the PAH concentrations observed in this study 

probably represent the peak in an annual cycle. 

Biological Response—Infauna 

A total of 11,505 invertebrates were identified in 18 sediment 

samples covering 0.0309 m2

 each collected in the vicinity of 

Bridge 148. These samples contained 69 taxa. Taxonomic 

codes used in the database are defined in Appendix C. 

Dominant taxa were defined as those representing 1% of the 

total infaunal abundance (count = 115). As with Bridge 146, 

the invertebrate community was dominated by annelids 

(ANNE = 2,383) and chironomids in the genus Chironomus 

(AICHSS = 7,396). In addition, mayflies in the genus 

Caenis (AICAE = 324), dipterans in the genus Bezzia

(AIBEZ = 267), and chironomids in the genus Dicrotendipes

(AIDTS = 145) also met the criteria. Dominant and subdominant taxa are listed in Table 15. These 19 taxa accounted for 99% of the animals recovered at Bridge 148. 

In Figure 17, total abundance, subdominant abundance, and 

the number of taxa observed in 0.0309-m2

 samples are compared with sediment PAH concentrations. Both measures of 

abundance and the number of taxa are depressed under the 

bridge and at a distance of 20 m downstream. It also appears 

that there is a mixed correlation between biological endpoints 

and the concentration of PAH in sediments. PAH levels are 

low and biological endpoints high at upstream and downstream stations. Between the bridge and the 6-m (20-ft) 

downstream stations, it appears that biological endpoints are 

positively correlated with sediment PAH concentrations. 

In Figure 18, Shannon’s and Pielou’s Indexes are compared 

with sediment concentrations of total PAHs. The values of 

these endpoints are low, suggesting a community dominated 

by a few opportunistic species, as has already been noted. 

The infaunal community has even lower diversity from under 

the bridge to a distance of 6 m (20 ft) downstream. The 

mixed response just noted suggests that there may be other 

Figure 16—Differences in upstream and 1.8 m (6 ft) 

downstream concentrations of various PAHs in 

sediments associated with 17-year-old Bridge 146 and 

2-year-old Bridge 148 crossing Pipe Creek in Cass 

County, Indiana. Naphthalene Acenaphthylene Acenaphthene

Fluorene

Phenanthrene

Anthracene

Fluoranthene

Pyrene

Benzo(a)anthracene

Chrysene

Benzo(b)fluoranthene

Benzo(k)fluoranthene

Benzo(a)pyrene

Benzo(ghi)perylene

Ideno(1,2,3-cd)pyrene

Dibenzo(a,h)anthracene

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

1.8

PAH at 6 ft downstream— 

PAH in upstream control

Bridge 146

Bridge 148

Compound 
34 

factors influencing the invertebrate community in this area of 

Pipe Creek. These factors were investigated through correlation analysis in Table 16 and the principle components analysis described in Figure 19. Total volatile solids and percentile 

fines (silt and clay) were included in the analysis but excluded from Table 16 because significant correlations were 

not observed. 

Sediment concentrations of both total PAHs and phenanthrene were at or below the TEL defined by Swartz (1999). 

However, 12 of the 14 biological endpoints were negatively 

correlated with phenanthrene, and five of those were significantly negative at α = 0.05 (Table 17). Only chironomids and 

specifically those in the genus Eukiefferiella sp. were significantly positively correlated with PAH. All other taxa were 

negatively correlated. That correlation suggests that either 

this robust community was moderately sensitive to PAH, 

particularly phenanthrene, or that sediment PAH concentrations were correlated with some other environmental factor 

responsible for the observations. 

Table 15—Dominant and subdominant invertebrate taxa 

observed in sediment samples collected in the vicinity 

of Bridge 148 crossing Pipe Creek in Cass County, 

Indianaa

 

Dominant taxa Code 

Found in 

number of 

samples 

Total 

abundance 

Annelids ANNE 18 2,383 

Chironomus sp. AICHSS 18 7,396 

Caenis sp. (Ephemeroptera) AICAE 16 324 

cf. Probezzia sp.(Diptera) AIPBEZ 8 267 

Dicrotendipes sp.(Diptera) AIDTS 16 145 

Sub-dominant taxa 

Gastropods 

 

Pleurocera sp. MGPLS 15 55 

Stagnicola sp. MGSTS 9 93 

Physella sp. MGPHLS 10 96 

Bivalvia 

Sphaerium sp. MPSPH 10 43 

Order Hemiptera 

Family Psyllidae AIPSY 9 70 

Order Coleoptera 

Oreodytes sp. AIORES 6 22 

Berosus sp. AIBERS 6 16 

Dubiraphia sp. AIDUBS 12 65 

Order Diptera 

Cryptolabis sp. AICRLS 3 20 

cf. Bezzia sp. AICBEZ 4 21 

Family Chironomidae AICHR 14 112 

Procladius sp. AIPRS 10 58 

Eukiefferiella sp. AEUKS 12 59 

Micropsectra sp. AIMPS 15 95 

Total abundance of dominants and subdominant taxa 11,340 

a

Eighteen samples, each covering 0.0309 m2, were collected. 

Figure 17—Number of taxa, total abundance, subdominant species abundance, and sediment concentrations of total PAH (TPAH) (micrograms TPAH per 

gram dry sediment) observed in the vicinity of 

Bridge 148 crossing Pipe Creek in Cass County, 

Indiana, November 4, 1997. 

Figure 18—Comparison of Shannon’s and Pielou’s 

Indexes with TPAH observed in sediments at 

Bridge 148 crossing Pipe Creek in Cass County, 

Indiana, November 4, 1997. 

 1200

0

200

400

600

800

1000

- 9.9 0 0.45 0.9 1.8 3.0 6.0 9.9

Distance from Bridge 148 (m)

Abundance and subdominant abundance

0

5

10

15

20

25

Number of taxa and total PAH

(μg PAH/g dry sediment)

Abundance Subdominant abundance Number taxa

TPAH 

0

0.5

1.0

1.5

2.0

2.5

-9.9 0 0.45 0.9 1.8 3.0 6.0 9.9

Distance from Bridge 148 (m)

Shannon's Index and Pielou's Index and 

TPAH (

μg TPAH/g dry sediment)

Shannon’s Index

Pielou’s Index

TPAH
35 

Figure 19—Principal components analysis of log(n + 1) transformed biological 

response variables and log transformed PAH concentrations in samples 

collected from Pipe Creek in the vicinity of Bridge 148 in Cass County, Indiana, 

November 4, 1997. 

Table 16—Pearson correlation matrix, describing the relationship between selected biological 

endpoints and independent factors that may have affected the invertebrate communitya 

Dependent variable Code Distance Depth RPD LTPAH Lphenanthrene 

Annelids ANNE 0.28 0.38 −0.01 −0.05 −0.29 

Pleurocera sp. LMGPLS 0.80 −0.76 −0.74 −0.60 −0.50 

Physella sp. LMGPHLS 0.86 −0.67 −0.75 −0.62 −0.60 

Caenis sp. LAICAE 0.90 −0.59 −0.88 −0.52 −-0.65 

Dubiraphia sp. LAIDUBS 0.22 0.40 −0.15 0.12 −0.32 

Bezzia sp. LAIPBEZ −0.58 −0.35 −0.71 −0.64 −0.85 

Family Chironomidae LAICHR −0.67 0.57 0.49 0.71 0.57 

Chironomus sp. LAICHSS 0.44 −0.27 −0.51 0.07 −0.08 

Dicrotendipes sp. LAIDTS −0.36 0.28 0.32 −0.06 −0.14 

Eukiefferiella sp. LAEUKS −0.65 0.59 0.66 0.71 0.60 

Total number taxa LTAXA 0.56 −0.26 −0.63 −0.13 −0.37 

Total abundance LABUNDAN 0.58 −0.22 −0.58 −0.03 −0.25 

Shannon’s Index LSHANNON 0.73 0.23 0.56 −0.63 −0.79 

Subdominant abundance LDABU148 0.96 −0.52 −0.87 −0.61 −0.78 

a

Significant (α = 0.05) correlations are in bold type. 

Factor 1

Factor 2

RPD 

SEDTVS 

FINES 

LANNE 

LMGPLS 

LMGPHLS 

LAICAE 

LAIDUBS 

LAIPBEZ 

LAICHR 

LAIDTS LAEUKS 

LTAXA 

LSHANNON 

LPHENAN LACENAPH 

LDABU148 

-1.0

-0.6

-0.2

0.2

0.6

1.0

-0.8 -0.6 -0.4 -0.2 0 0.2 0.4 0.6 0.8 1.0

LABUNDAN 
36 

The Bridge 148 sample transect was chosen because of its 

slow currents and depositional characteristics that would 

allow for the accumulation of PAHs lost from the bridge. 

Large quantities of maple leaves (Acer sp.) had accumulated 

along the sample transect, particularly between the downstream perimeter of the bridge and the 6-m (20-ft) downstream sample station (Fig. 14). 

The reduction–oxidation potential discontinuity (PRD) is an 

indication of the depth to which sediments are aerobic. At 

Bridge 148 in the fall of 1997, the depth of the RPD appeared most influenced by the accumulation of maple leaves. 

These leaves had recently fallen and had not yet started to 

decompose. However, numerous aquatic insects were found 

in the leaf mats (as expected). Note in Table 16 that the depth 

of the RPD is negatively correlated with the abundance of 

most taxa. This indicates that where an abundance of organic 

material (maple leaves) was found, the RPD was reduced, but 

the abundance of most taxa was increased. As shown in 

Figure 19, the depth of the RPD and sediment concentrations 

of phenanthrene were positively correlated (r = 0.52). This 

suggested that greater sediment concentrations of phenanthrene were found in areas where the RPD was deeper and 

the leaf mat was reduced. Leaf matting was not quantified 

during the survey, and no direct measure of this factor was 

possible. Therefore, this discussion remains purely hypothetical. The value of the triad approach (physicochemical 

analysis + invertebrate community analysis + laboratory 

bioassay analyses) to environmental risk assessment is 

clearly evident in this instance. 

Biological Response—Sediment Bioassays 

Results of 10-day sediment bioassays using the amphipod 

Hyalella azteca are presented in Table 17. Significant 

differences were not observed in the 10-day survival of test 

animals between stations. Survival at downstream stations 

was equal to or greater than survival in laboratory test sediments and greater than survival in sediments collected at the 

upstream control. Results from these bioassays suggest that 

sediment toxicity was not responsible for the differences 

observed in the resident invertebrate community. 

Risk Assessment Summary of Bridges 146 and 148 

Results of these surveys are consistent with our existing 

knowledge regarding the physics and chemistry of PAHs, the 

loss of creosote from pressure-treated wood, and the biological response to sedimented PAHs. The following summarizes 

our findings: 

• Creosote does bleed from pressure-treated wood. In freshwater, at 17o

C, the rate is approximately 37 μg/cm2-day. 

For a Class A piling that is 30 cm in diameter and submerged in 60 cm of water, the total surface area is 

5,652 cm2

. The piling loses 0.213 g of PAHs per day during the first year, equivalent to a drop of crankcase oil. 

That amount decreases exponentially with time and by the 

mid-point of piling life (about 35 years), PAH loss is reduced to about 3% of these initial losses. This temporal 

reduction was evident in the Pipe Creek data. Sediment 

concentrations of PAH downstream from the 17-year-old 

Bridge146 were much less than those observed at the 

2-year-old Bridge 148. 

• PAHs observed in the vicinity of both Bridges 146 and 148 

were rich in acenaphthylene, fluorene, phenanthrene, and 

fluoranthene. These intermediate weight compounds are 

consistent with new creosote contamination, and the bridge 

structure appears to be the source. As sedimented creosote 

ages, the low and intermediate weight compounds are metabolized by microbes, leaving a deposit rich in the highmolecular-weight compounds. This shift in PAH species 

was not observed at these two bridges, despite the 15-year 

difference in their ages. High summer temperatures likely 

exacerbate the loss of creosote oil from the elevated portions of the bridge. These compounds accumulate in sediments during low summer flows and likely reach a peak 

during the late summer and fall (when this survey was 

completed). Before these accumulations can weather (as 

indicated by a preferential loss of the intermediate weight 

compounds), the deposits are diluted and redistributed 

downstream during high winter and spring flows. 

• Elevated sediment concentrations of PAH were observed 

downstream from both bridges. Concentrations of naphthalene, acenaphthylene, and phenanthrene exceeded the PEL 

defined by Swartz (1999), and adverse biological effects 

could have been expected in very sensitive taxa at the new 

Bridge 148 but not at the older Bridge 146. 

• Sediment concentrations of PAH downstream from Bridge 

146 were low, and acenaphthene and phenanthrene barely 

exceeded the TEL defined by Swartz (1999). Adverse effects were not expected in the invertebrate community, but 

Table 17—Summary results of Bridge 148 sediment 

bioassays using the test amphipod Hyalella azteca

 Percentage of surviving amphipods with an initial 

count of 20 per replicate 

Replicate 

Laboratory 

control 

Upstream 

control 

+0.45 m 

(+1.5 ft) 

+1.8 m 

(+6.0 ft) 

+9.9 m 

(+33.0 ft) 

1 100 100 90 100 100 

2 100 90 100 100 100 

3 90 100 100 100 100 

4 100 100 100 100 100 

5 100 100 100 100 100 

6 100 90 100 100 100 

7 100 100 100 100 100 

8 100 100 100 100 100 

Mean 98.8 97.5 98.8 100 100 
37 

significant reductions in most biological endpoints were 

observed between the downstream perimeter of the bridge 

and 6 m (20 ft) downstream. In contrast, no sediment toxicity was observed in laboratory bioassays when survival 

in downstream sediments was compared with either laboratory or upstream control sediments. These results and the 

analysis suggest that the invertebrate community was more 

affected by the presence of various degrees of coarse particulate organic matter (CPOM) in the form of heavy mats 

of recently dropped maple leaves than on the low concentrations of PAH found in adjacent sediments. 

• Adverse effects were not observed in either the resident 

invertebrate community at Bridge 146 or laboratory bioassays. Pipe Creek was chosen for this assessment because of 

the slow current speeds that would allow PAHs lost from 

the bridge to accumulate in the immediate vicinity of the 

bridge. The landscape surrounding Pipe Creek is devoted 

to agriculture (primarily corn), which likely contributes to 

the significant sand, silt, and clay bedload carried by the 

stream. The combination of slow currents and fine-grained 

sediments results in a stressful environment that does not 

support a sensitive invertebrate community. The invertebrate community in Pipe Creek is dominated by annelids 

and chironomids, both considered robust taxa. That is 

likely the reason that adverse effects were not observed 

when they could have been expected in a more sensitive 

community. It could be argued that a similar bridge placed 

over a faster flowing stream with a more diverse and stable 

benthic habitat (more cobble and boulder with less sand) 

would result in the observation of adverse biological effects. However, habitats characteristic of more sensitive 

communities with large numbers of EPT taxa include 

higher current speeds. These higher current speeds would 

result in significantly reduced concentrations of PAH. 

Modeling of PAH concentration as a function of current 

speed suggests that in water bodies where current speeds 

exceed 25 to 30 cm/s, PAHs would be dispersed to the 

point of not being detectable in sediments. 

• The point is that sediment quality criteria are frequently 

based on the most sensitive taxa. These criteria do not take 

into account the relationships between water body physicochemical characteristics (for example, current speeds), 

community composition as a function of those physicochemical characteristics, and the transport and deposition 

of potential contaminants as a function of those same factors. These results are similar to those observed by Brooks 

(2000) in an examination of the biological response to 

handicapped access boardwalk construction in a series of 

beaver ponds in the Wildwood Recreational Area of 

Oregon. 

• Both of these creosote-treated bridges appeared relatively 

clean and dry with little evidence of preservative weeping 

from overhead components. However, preservative does 

migrate out and run down the surface of piling (Fig. 20). 

It is also likely that some preservative drips from overhead 

structures during hot summer days. Oregon State University is evaluating losses from overhead structures as a function of preservative retention and ambient temperature. 

That work will significantly improve our knowledge of 

creosote loss rates from treated bridges and its transport in 

aquatic environments. 

Penta-Treated Bridges 

Penta-treated wood was identified during the site selection 

survey as the second most commonly used preservative in the 

construction of timber bridges. 

Sources 

The U.S. Department of Agriculture (USDA 1980) noted that 

penta is “ubiquitous in aquatic environments and its sources 

are unclear.” Observable levels of penta can result from 

direct contamination, degradation of other organic compounds, or chlorinated drinking water. The USDA report 

notes that circumstantial evidence, including the detection of 

penta in rainwater, indicates that penta may occasionally be 

present in ambient air. 

Historically, penta has been used extensively in agriculture 

and industry as an insecticide, fungicide, herbicide, algaecide, and disinfectant. However, the major commercial application of technical grade penta is in the wood preservation 

industry. Eisler (1989) reported that 80% of the 23 × 106

 kg 

of technical penta produced annually is used for wood preservation. He noted that penta is found at levels from <1 to 

7.3 μg/L in some British Columbia waters. In general, 

Eisler’s (1989) data imply a correlation between industrialurban centers and increasing penta concentrations in the 

water column. 

Figure 20—Creosote weeping from pressure-treated 

piling at Bridge 148. 
38 

Matsumoto (1982) noted that vascular plants and their detritus naturally produce many phenolic acids. However, he 

stipulated that he found no evidence of naturally occurring 

penta in his study of the polluted Tokyo River water and 

pristine river or reservoir and pond waters on the Bonin 

Islands. No information was obtained that suggested that 

penta is biosynthesized, and there are no known significant 

natural sources of penta. However, Lampi and others (1992) 

observed up to 30 to 70 μg penta/kg dry sediment in cores 

dating back to the 17th century. They concluded that this may 

have been due to wood burning and aerial transport, because 

the lake is about 8 km (5 miles) from the nearest settlement 

and tens of kilometers from the nearest industrial center. 

Environmental Chemistry 

Penta (C6Cl5OH) has a relative molecular mass of 266.34, a 

melting point of 190o

C, a boiling point of 310oC, and a density of 1.98 g/cm3

 at 22oC. The vapor pressure is 0.00415 Pa 

at 20o

C. Penta has an octanol–water partition coefficient (log 

Kow) of 3.3 at neutral pH, 5.1 at pH 4, and 1.9 at pH 8. Penta 

readily dissolves in most organic solvents. Its solubility in 

water is also pH dependent and varies between 10 mg/L at 

pH 6 to 20 mg/L at pH 8 (Mackay and others 1995). 

In the 1970s and 1980s, scientific recognition of the toxicity 

of some dioxins led to increased public concern. The dioxin 

2,3,7,8-tetrachlorodibenzo, commonly called 2,3,7,8-TCDD, 

was recognized as an exceptionally toxic compound, at least 

to some test organisms. This fact, coupled with the knowledge that this compound was present in Agent Orange (the 

defoliant used in Vietnam), heightened concern. Eisler 

(1989) noted that many commercial samples of technical 

grade penta were heavily contaminated with a large number 

of potentially toxic compounds including dibenzofurans, 

dioxins, and hexachlorobenzenes. The relative toxicity and 

accumulation potential of some of these contaminants exceeded that associated with the parent penta by several orders 

of magnitude (Huckins and Petty 1981). For example, Eisler 

(1986) reported that some isomers of hexachlorodibenzodioxin, present in technical grade penta at concentrations of 

1,000 to 17,300 μg/kg during the 1970s were lethal to guinea 

pigs at doses of 60 to 100 μg/kg body weight. The dioxins in 

penta of greatest concern are the hexachlorodibenzo-pdioxins (HxCDD). The U.S. EPA has limited the concentration of this compound to 2 mg/kg in commercial penta sold 

in the United States. A maximum concentration in any one 

batch of 4 mg/kg is allowed. 

Penta is not dissociated from its hydroxyl ion in aqueous 

solutions at pH < 5.0. However, as the pH increases, the 

bioavailability and toxicity of penta decrease because less of 

the compound is found in the undissociated form. The solubility of penta and potential for adsorption to suspended 

inorganic particulate matter (particularly clay) is positively 

correlated with the dissociated fraction, which increases 

significantly at pH > 6.5. 

Fate 

Penta may be dissolved in water or sorbed to suspended 

matter or bottom sediments. In addition, penta is taken up by 

fauna and flora that metabolize the compound at various 

rates. Penta readily degrades in the environment by chemical, 

microbiological, and photochemical processes (USDA 1980, 

Eisler 1989). Photochemical degradation is a function of the 

spectrum and intensity of incident light and appears to proceed rapidly in natural environments with half-lives of 0.15 

to 15 days (Smith and others 1987). The degradation of penta 

in sediments depends on a number of environmental factors 

that are discussed in the following sections. Half-life in 

sediments can range from days to years, depending on environmental conditions. The ultimate fate of penta appears to 

be burial in anaerobic sediments under infrequently encountered conditions or mineralization to carbon dioxide and 

water. 

Dissolved Penta 

Penta that is dissolved in water may be removed by volatilization, photodegradation, absorption, or biodegradation. 

Penta is subject to rapid photodegradation under laboratory 

conditions. Boyle and others (1980) examined the degradation of penta in a two- by two-array of microcosms that did 

and did not contain natural lake sediments held under aerobic 

and anaerobic conditions. At the end of the 131-day experiment, the authors determined the penta half-life under each 

of the four conditions. Their results are summarized in 

Table 18. Fisher (1990) concluded that in aerobic and organically rich environments, the half-life of dissolved and 

sedimented penta would be about 1 week. Middaugh and 

others (1993) determined that the gram-negative bacterium 

Pseudomonas sp. (strain SR3) was able to degrade penta and 

that it provided an adequate sole carbon source sustaining 

growth. Nearly complete degradation of 39,000 to 40,000 μg 

Table 18—Pentachlorophenol remaining in water and 

sediments at the end of 131 days in each of four 

aquariaa

 

Test conditions 

Total 

penta in 

water 

(mg) 

Water 

halflife 

(days) 

Penta in 

sediments 

(mg) 

Total 

penta 

(mg) 

Aerobic without 

mud (lighted) 

0.95 18.6 0.950 

Aerobic with mud 

(lighted) 

0.21 13.9 0.03 0.240 

Anaerobic without mud (dark) 16.00 79.8 16.000 

Anaerobic with 

mud (dark) 

0.005 12.8 0.04 0.045 

a

Penta (100 mg) was originally added to each of the 

 aquaria. Water column half-life is given for penta as 

 determined in this experiment. 
39 

penta/L was accomplished by acclimated Pseudomonas sp. 

Penta half-lives in freshwater streams reported in McAllister 

and others (1996) varied between 40 and 120 h. 

Boyle and others (1980) also partitioned 14C at the end of the 

experiment. The bulk (99%) of the sedimented 14C was observed in the non-biogenic clay fraction. They also found 14C 

in algae, floating flocculent material, and other biogenic 

material in the water column. Minimal 14C was observed 

elsewhere in the microcosms (including the aquarium sides 

and cover). The authors concluded that penta degradation 

was positively correlated with incident light levels, pH, 

oxygen, and the presence of sediment. They concluded that 

penta is likely most persistent in the deoxygenated 

hypolimnion water of lakes. Several environmental factors 

affect the rate at which penta is degraded in natural aquatic 

environments. 

Effects of pH and Water Temperature 

on the Degradation of Penta 

Valo and others (1985) found that the metabolism of penta 

was inhibited at less than 8o

C or greater than 50oC. Optimum 

degradation occurred at 28o

C. Jarvinen and Puhakka (1994) 

and Jarvinen and others (1994) found that 99% of the penta 

present in contaminated groundwater was degraded at 5o

C to 

10o

C. Trevors (1982) documented no penta degradation by 

acclimated Pseudomonas at 0o

C. Penta degradation rates at 

4o

C were dependent of the specific Pseudomonas strain used. 

However, an average of 28.2% of the initial 50,000 μg 

penta/L substrate was metabolized in 80 days at 4o

C, and 

50.2% of the same concentration was metabolized in 8 days 

at 20o

C. 

Valo and others (1985) observed penta degradation at pH 

values between 5.6 and 8.0. A neutral or slightly acidic pH 

was found to be optimum. Wong and Crosby (1981) observed penta half-lives of approximately 100 h at pH 3.3 and 

3.5 h at pH 7.3 in sterile solutions containing 100,000 μg 

penta/L. Penta degradation was not observed in flasks maintained in the dark. 

It appears that penta degradation is optimum at pH values 

between 6.5 and 8.0 and at temperatures between 10o

C and 

30o

C. These are conditions expected in much of North America during all seasons except winter, when low temperatures 

at northern latitudes can be expected to decrease penta degradation rates. In addition, penta is expected to degrade more 

slowly in areas subjected to low pH.

Microbial Community Adaptation 

Larsson and others (1993) and Larsson and Lemkemeier 

(1989) observed significantly higher penta degradation by 

unacclimated microbial communities inhabiting brown water 

lakes containing high levels of humic acid when compared 

with clear water lakes. These authors concluded that the 

microbe communities inhabiting brown water lakes had 

adapted to the higher phenol levels naturally present in the 

water and therefore were pre-acclimated to metabolize penta. 

McAllister and others (1996) reviewed the literature pertaining to the microbial degradation of penta. They confirmed 

that sediment penta levels exceeding 300 μg/kg inhibit microbial degradation until a period of acclimation has passed. 

When acclimation began, it appears that the higher the initial 

concentration of PCP, the longer the maximum number of 

viable cells, capable of degrading PCP, was maintained. 

Gonzalez and Hu (1991) observed that lag phases of 10 h 

occurred at 10 mg/L, 30 h at 20 mg/L, 55 h at 44 mg/L, 80 h 

at 80 mg/L, and 200 h at 200 mg/L. 

In summary, it appeared that microbial communities were not 

generally preadapted to metabolize penta. Initial exposure of 

naïve communities to penta concentrations as low as 300 μg 

penta/L can reduce growth. An adaptation of several hours 

to perhaps 2 weeks is necessary for community adaptation 

to penta. Following this time, adapted communities can 

tolerate much higher concentrations of penta and rapidly 

metabolize it. 

Effects of Additional Carbon 

Sources on Penta Metabolism 

Topp and others (1988) studied the response of pentadegrading Flavobacterium sp. to high levels of penta with 

and without the addition of sodium glutamate as a cometabolite. They found that the specific activity of penta-degrading 

cells in the absence of supplementary carbon was 

1.51 × 10−13 g penta/cell-h. They showed that the form and 

amount of alternative substrates were important in determining the metabolism of penta. For instance, optimal stimulation of penta removal required the addition of 3.0 g sodium 

glutamate/L. However, glutamate in combination with glucose or cellobiose partially repressed pentametabolism. 

Flavobacterium removed 2.5% of the penta from a 25,000 

μg/L initial concentration. The addition of 4 g sodium glutamate/L increased metabolism, resulting in the removal of 

61.9% of the penta in 3 h. However, when the mixture was 

amended with 4 g sodium glutamate and 5 g/L glucose, penta 

metabolism was reduced to 15.5% in 3 h. It was the combination of the two substrates that reduced penta metabolism, 

because in a separate experiment the authors found that the 

addition of 0.5 g of glucose to the medium in the absence of 

sodium glutamate resulted in the complete degradation of an 

initial 61,000 μg penta/L solution by Flavobacterium sp. The 

amount of penta removed decreased when incremental 

amounts of sodium glutamate were added to the glucose. 

Topp and others (1988) did not discuss the possibility that 

Flavobacterium sp. acclimated and degraded penta when it 

represented a sole carbon source but preferentially shifted to 

alternate substrates when available. Perhaps rather than 

acting antagonistically, the combination of sodium glutamate 

and glucose acted synergistically, reducing the dependence 

of the bacteria on penta. Yu and Ward (1994) observed 
40 

maximum penta degradation in medium supplemented with 

glucose and peptone. 

Topp and others (1988) found that amendment with supplementary source of carbon reduced the lag time required 

before significant penta metabolism commenced. These 

authors noted that penta concentrations greater than 

20,000 μg/L inhibited Escherichia coli and that Pseudomonas was inhibited at concentrations greater than 500,000 

μg/L. They also reported that 50,000 μg penta/L was degraded as a sole source of carbon after a lag phase of 90 h. 

Penta half-lives in freshwater streams reported in McAllister 

and others (1996) varied between 40 and 120 h. Consistent 

with Liu and others (1981), McAllister and others (1996) 

noted that additional substrates tend to reduce penta degradation rates and hypothesized that adsorbed penta is less 

bioavailable. McAllister and others (1996) also noted that the 

most widely studied penta degrading microorganisms are the 

pure culture bacterial strains, Flavobacterium and 

Rhodococcus chlorophenolicus. The enzymes responsible for 

initiating the catabolism of penta by Flavobacterium sp. have 

been isolated and characterized. Furthermore, the genes 

encoding these enzymes have been characterized and cloned 

into E. coli, which then demonstrated the ability to 

degrade penta. 

These reports suggest that penta will be degraded more 

rapidly in organically rich environments. However, some 

caution is necessary because there is evidence that some 

combinations of organic substrates appear to result in slightly 

reduced degradation rates. 

Effects of Water Hardness on Dissolved Penta 

Brockway and others (1984) studied the fate of penta in static 

and continuous-flow hard and soft water mesocosms. They 

observed no significant effect on the fate or effects of penta 

associated with water hardness. 

Sedimentation 

Penta is moderately persistent in soil. Published data indicate 

that penta can persist in soils for various times, ranging from 

weeks (21 days) to 5 years. Under most conditions, penta 

will seldom persist in the soil for more than 9 months and its 

half-life will frequently be far less. Numerous studies have 

identified soil microorganisms capable of penta degradation. 

In most studies of penta biodegradation, acclimated populations of microorganisms have been utilized. The current 

major use of penta is for wood preservation; therefore, the 

likeliest source of soil contamination is leaching or bleeding 

of the preservative from treated wood. Such phenomena may 

result in low levels of penta contamination within meters of a 

treated pole. 

Fisher (1990) constructed microcosms with water at pH 4, 6, 

and 8 and sediments with 0.0% and 3.0% total organic 

carbon (TOC). She found that more penta was partitioned to 

the sediments at lower pH than at higher pH values. In addition, she reported a positive correlation between sediment 

TOC and penta concentration. The organisms in the high 

TOC microcosms accumulated significantly less penta than 

did those in the 0.0% TOC systems. This work is consistent 

with Eisler’s (1989) observation that at low pH, penta is fully 

protonated and lipophilic, whereas at high pH, it is ionized 

and unlikely to adsorb to organic ligands. 

Shimizu and others (1992) determined the adsorption coefficients of penta in aquatic environments with various organic 

carbon contents (0.72% to 2.38%) and clay contents (10.1% 

to 60.8%). They concluded that at pH values between 6 and 

8, the adsorption coefficient was not influenced significantly 

by organic carbon content (correlation coefficient = 0.12) but 

was positively correlated with clay content (correlation coefficient = 0.94). This work suggests that clay particles (which 

frequently carry an electrical charge), rather than particulate 

or dissolved organic carbon, form a more likely adsorption 

nucleus in aquatic environments. 

In summary, it appears that the potential for sedimentation of 

penta is a complex problem driven by at least the following 

parameters: 

• At reduced pH values, penta tends to be fully protonated 

and more lipophilic. Therefore, it would be expected to 

adsorb to dissolved or particulate organic matter. In contrast, at higher values of pH, more penta is expected to be 

ionized, with a higher potential for binding to polar 

adsorption nuclei represented by particulate inorganic 

matter (silt and clay). 

• It appears that more penta will be partitioned from the 

water column to sediments having increased organic carbon content. This can have a major effect in removing 

penta from the water column. 

Degradation of Sedimented Penta 

Bryant and Rogers (1990) described the degradation of penta 

in anaerobic sediments from diverse locations around the 

world. They observed that dechlorination did not occur for at 

least the first 15 days of exposure in unadapted sediments. 

However, following that adaptation, penta was completely 

degraded by 33 days. A second addition of 70 mg penta/kg 

sediment on day 33 was rapidly dechlorinated to about 

25 mg/kg in 2 days. In contrast, they observed no biotransformation within 40 days after penta was added to unadapted 

Cherokee Pond sediments. The point made in this study is 

that not all microbes have the ability to dechlorinate penta as 

a first degradative step. Their study is consistent with other 

reports, indicating that unacclimated, but biological rich, 

sediments require approximately 2 weeks for development of 

suitable microbial communities before aerobic or anaerobic 

degradation of penta begins. However, when established, 
41 

these communities rapidly catabolize penta. Interestingly, 

these authors found no degradation of penta in autoclaved 

sediments, further emphasizing the microbial nature of the 

observed degradation. 

Van Gestel and Ma (1988) determined penta half-lives in low 

pH and organically rich Holten (pH 5.6; 6.1% organic matter) and Kooyenburg (pH ~5.0; 3.7% organic matter) soils of 

23.2 to 47.9 days, respectively. There was no significant 

difference in the half-lives of penta in these two sediments. 

Smith and Novak (1987) found that penta concentrations as 

high as 25,000 μg/L in saturated soils were degraded to 

nondetectable levels in less than 3 months. They found that 

chlorophenol degradation rates were linearly related to the 

initial concentration and varied between 100 μg/L-day at 

200 μg/L initial concentration to greater than 10,000 μg/Lday at an initial concentration of about 800,000 μg/L. 

Effects of Sediment Oxidation–Reduction 

on Penta Degradation 

Delaune and others (1983) examined the degradation of 

penta in estuarine sediments following a major accidental 

spill in a Louisiana Gulf Coast estuary. They completed a 

series of laboratory experiments in which pH was manipulated between 5.0 and 9.0 and sediment redox between –250 

and +500 mV. The authors found maximum degradation at 

pH 8.0, with declines at either lower or higher values. Penta 

was observed to degrade at all values of redox. However, 

significantly higher degradation rates were observed under 

aerobic (+250 and +500 mV) than with reducing (0.0 and 

–250 mV) conditions. A half-life of about 24 days was apparent at pH 6.5 and a redox potential of +500 mV. However, at this pH, minimal degradation was observed at any 

other value of redox. At pH 8.0, significant degradation was 

observed at +250 and +500 mV redox potentials and a halflife of 26.5 days was apparent at +500 mV. The authors 

found that penta was more tightly bound to oxidized sediment solids than to reduced sediments. Therefore, they concluded that there was a tendency for penta to become preferentially associated with the thin oxidized surface sediment 

horizon as well as with suspended colloidal particulates, 

which would also tend to be oxidized. Under either condition, penta would be retained in the photic zone of shallow 

estuaries where the potential for photodegradation would be 

enhanced. The authors concluded that although tidal transport and photodegradation in the water column could play a 

role in the removal of residual penta from a spill area, their 

laboratory studies suggested that degradation under either 

aerobic or anaerobic conditions could account for the disappearance of residual penta left in the immediate vicinity of 

the spill or that which was transported from the spill site and 

deposited onto the bottom in adjacent water bodies. 

Guthrie and others (1984) studied the anaerobic degradation 

of penta as a component of sewage sludge during treatment. 

They found that methanogenic bacteria were unaffected by 

penta concentrations less than 200 μg/L. Acclimation of the 

bacterial flora to penta increased the inhibition threshold to 

about 600 μg/L. They found that penta was biodegradable 

anaerobically and that removal was so complete that the 

soluble concentrations in the Phase II reactors were below 

detection limits of 5 μg/L. Sorption appeared a minor 

mechanism of penta removal, and volatilization was considered insignificant. The authors concluded that penta undergoes extensive anaerobic biodegradation, especially by acclimated microbial communities. Anaerobic degradation of 

penta was confirmed by Kudo (1989). 

Liu and others (1981) observed similar results in their comparison of aerobic and anaerobic degradation of sedimented 

penta at pH 7.0. They observed an increase in the aerobic 

half-life of 0.36 days to 190 days in anaerobic conditions. 

They also found that the inclusion of either sodium chlorophenate or glucose as a second substrate inhibited rather than 

enhanced the anaerobic degradation of penta. In contrast, 

Valo and others (1984) found that penta degradation was 

enhanced by the addition of 0.4 or 40 mM ammonium chloride (NH4Cl). 

Bryant and Rogers (1990) examined the degradation of penta 

in anaerobic sediments from Georgia, Florida, New York, 

and the Soviet Union. They observed an adaptation of about 

15 to 20 days during which little penta degradation occurred. 

Following microbial adaptation, degradation in anaerobic 

sediments was rapid with an unstated half-life of 2 to 7 days. 

Penta half-lives taken from the literature are summarized in 

Table 19. 

Data provided in Baker and others (1980), DeLaune and 

others (1983), and Boyle and others (1980) were interpreted 

to produce half-lives as a function of initial penta concentration, reduction oxidation potential, pH, and temperature. 

Sedimented half-life was estimated assuming that microbes 

required 15 days to adapt to penta and that degradation rates 

remained linear at all times. This is a small database, and the 

methodology is not precise. However, the results are consistent with the remainder of the literature and appear to reasonably predict sedimented penta half-lives that are important in understanding its accumulation in sediments. Data 

from these studies are summarized in Table 20. 

These data were analyzed using linear and nonlinear regression analyses. Within the ranges of data at hand, sedimented 

penta half-life was not a function of the initial concentration 

(p = 0.10) or temperature (p = 0.40). In addition, the constant 

term was not significant (p = 0.42). Redox potential and pH 

were significant factors (p = 0.001 and p < 0.000, respectively). The final regression was highly significant 

(p < 0.00015), and it explained 76% of the variation in the 

database. The underlying assumptions requiring normally 
42 

distributed residuals and homoscedasticity were met. The 

resulting predictive equation is 

Sedimented penta half-life 

 = 18.19448 × pH – 0.29284 × redox (mV) (5) 

At pH 7.5, this equation predicts a penta half-life of 63 days 

in reasonably well-oxygenated (+250 mV) sediments. In 

reducing sediments (−100 mV), the half-life is increased to 

165.7 days. Figure 21 provides a detailed representation 

using a quadratic smoothing routine in the Statistica software 

package (StatSoft, Tulsa, Oklahoma). Under normal 

environmental conditions, the redox potential in surface 

sediments will vary between –100 and + 400 mV, and the pH 

is expected to vary between 6 and 8.5. Figure 1 suggests that 

under these conditions, sedimented penta half-lives will vary 

significantly between 12.7 and 176.4 days. The linear equation used to make predictions in this model predicts a halflife of 24.8 days at pH 7.8 and +400 mV and 156.6 days at 

pH 7.0 and –100 mV. 

A three-dimensional display of sedimented penta half-life as 

a function of redox and pH is shown in Figure 22. This figure 

was constructed using the distance weighted least squares 

routine in the Statistica software package. 

Table 19—Summary of pentachlorophenol half-lives and the effects of various environmental parameters on degradation rates 

Reported penta half-life in water Half-life (days) Conditions 

Boyle and others (1980) 18.6 Aerobic in the laboratory 

Boyle and others (1980) 79.8 Anaerobic in the laboratory 

Crossland and Wolff (1985) 2.0 to 4.7 Outdoor mesocosms 

Liu and others (1981) 0.36 Aerobic in the laboratory 

Liu and others (1981) 190.0 Anaerobic in laboratory 

Wong and Crosby (1981 2.0 pH 7.3 (natural sun-light) 

Yu and Ward (1994) ~1.5 Mixed bacterial cultures 

Effects of pH (7.3) on penta half-life Half-life (h) Conditions 

Wong and Crosby (1981) 3.5 Laboratory F40BL lamps 

Wong and Crosby (1981) 100.0 Laboratory F40 BL lamps 

Wong and Crosby (1981) 48.0 Natural sunlight 

Effects of ambient temperature on penta half-life Half-life (days) Conditions 

Topp and others (1988) 4.0 oC >80 Pseudomonas cultures 

Topp and others (1988) 20 oC <12 Pseudomonas cultures 

Anticipated half-life of penta in soils or sediments Half-life (days) Conditions 

DeLaune and others (1983) 24 to 26 Natural estuarine sediments 

McAllister and others (1996) 10 to 70 Flooded soils 

McAllister and others (1996) <2 to >5 Freshwater streams 

Neary and others (1990) Average 30 Southern ecosystems 

Smith and Novak (1987) <5 9 to 13 mg PCP/L unsaturated soil 

Smith and Novak (1987) ~15 Average 3 mg PCP/L unsaturated soil 

Effects of sediment reduction–oxidation potential Half-life aerobic (days) Half-life anaerobic (days) 

Boyle and others (1980) 13.9 12.8 

Bryant and Rogers (1990) 2 to 5 

McAllister and others (1996) 144 

Penta half-life in plants and animals Species Penta half-life (h) 

Benner and Tjeerdema (1993) Atherinops affinis 52.7 

Glickman and others (1977) Onchorhynchus mykiss 6.2 to 6.9 
43 

Figure 21—Quadratic solution to the penta half-lives in 

Table 3, with sediment reduction–oxidation potential 

(millivolts) and pH as independent factors. (z = −715.685 

+ 243.79x − 1.042y − 17.252x

2

 + 0.073xy + 0.001y

2

). 

Case Studies 

Seidler and others (1986) examined the transport and fate of 

penta-contaminated wastewater as it passed through two 

estuarine ponds being used to grow shrimp in Florida. Penta 

was added to the test ponds by broadcasting 500 mg of sodium-pentachlorophenate in acetone on day zero, followed 

by the addition of 250 mg on alternate days thereafter to 

provide a theoretical concentration of 10 μg penta/L. Penta 

concentrations of 3 to 5 μg/L were observed in the treatment 

Figure 22—Half-life of sedimented penta in days on the 

Z axis as a function of sediment reduction–oxidation 

potential and pH. The three-dimensional plot was 

developed using a distance-weighted least square 

routine. 

Table 20—Estimated sedimented pentachlorophenol half-lives as a function of initial 

penta concentration, reduction–oxidation potential, pH, and temperaturea 

Source 

Initial penta 

concentration 

(μg/kg) 

Redox 

potential 

(mV) pH 

Temperature 

(

o

C) 

Half-life 

(days) 

Baker and others 1980 100.0 250 7.1 0.0 47.5 

Baker and others 1980 100.0 250 6.9 20.0 36.7 

Boyle and others 1980 1,400.0 0 4.5 15.0 13.0 

Boyle and others 19080 1,400.0 250 6.8 15.0 7.5 

DeLaune and others 1983 20.0 500 6.5 33.1 6.0 

DeLaune and others 1983 20.0 250 6.5 33.1 103.5 

DeLaune and others 1983 20.0 0 6.5 33.1 178.8 

DeLaune and others 1983 20.0 500 8.0 33.1 12.0 

DeLaune and others 1983 20.0 250 8.0 33.1 32.5 

DeLaune and others 1983 20.0 0 8.0 33.1 81.0 

DeLaune and others 1983 20.0 −250.0 8.0 33.1 290.0 

DeLaune and others 1983 20.0 500.0 9.0 33.1 50.0 

DeLaune and others 1983 20.0 250.0 9.0 33.1 50.0 

a

Sedimented penta half-lives were estimated from data. 

 

0

pH

Redox potential (mV) -400

-200

200

400

600

4 5 6 7 8 9 10

12.7 days

45.5 days

45.5 days

12.7 days

78.2 days

110.9 days

143.6 days

176.3 days

209.1 days 241.8 days

274.5 days

307.3 days

Redox potential (mV)

 340 

 280

 220

 160

 100

 40

 10

9

8

7

6

5

4 800

400

200

0 –200

–400

pH

Half-days
44 

ponds during the first 22 days of the study. At pond pH 8.0, 

the authors predicted that 99.9% of the added penta would be 

dissolved in the phenolate form, suggesting that photolysis 

would be the most significant degradation process. They 

calculated a penta half-life of 2 days for the ponds being 

treated. 

Seidler and others (1986) claimed that sediments contained 

54 times more penta than was found in the water column. 

However, data presented in their paper suggest that sediment 

levels mimicked water column levels with a concentration 

factor of 1 to 10. Sediment concentrations of penta averaged 

about 5 μg/kg with a maximum of about 18 μg/kg. The authors concluded that their results suggested that a chronic 

influx of penta at a concentration of 10 μg/L to an estuarine 

environment resulted in elevated concentrations in the water, 

sediment, and shrimp. Penta concentrations in shrimp 

(maximum of about 18 μg/g on day 20) were positively 

correlated with water column concentrations rather than with 

sediment concentrations. This suggests direct uptake from the 

water rather than biomagnification from infaunal and epifaunal prey. When the source of penta was removed, the compound rapidly disappeared from the water, shrimp, and sediments (in a matter of hours or days). 

Crossland and Wolff (1985) determined the half-life of penta 

in outdoor ponds repeatedly dosed to maintain a concentration of 50 to 100 μg penta/L. The authors hypothesized that 

evaporation, sorption, hydrolysis, biodegradation, and indirect phototransformation of penta would be of minor importance under the environmental conditions at the ponds. The 

partition coefficient for penta between water and sediments 

was predicted to be near unity. The authors used the SOLAR 

mathematical model to calculate a direct photodegradation 

rate constant for the transformation of penta in the ponds 

where pH varied between 7.3 and 10.3, with a mean of 8.3. 

Several bioassays were conducted, and pond invertebrates 

were enumerated at levels of taxonomy exceeding Order. The 

results appeared consistent with the general body of literature 

describing the toxicity of penta. The observed half-lives of 

penta in the three treatment ponds varied between 2.0 and 

4.7 days and were in good agreement with the predicted halflives based on results of the SOLAR analysis. The authors 

concluded that direct phototransformation was responsible 

for nearly all the penta degradation observed in this study. In 

addition, they noted that at the end of the study, sediment 

concentrations of penta were very similar to water column 

concentrations. They hypothesized that insufficient time had 

elapsed for development of a microbial community capable 

of efficiently metabolizing penta in the sediments. 

Robinson–Wilson and others (1983) examined the degradation of penta in a series of experimental ponds contaminated 

by a single high dose of penta (1,000 μg/L) followed by a 

series of small doses (0.2, 0.2, 0.4, and 0.4 μg/L) at monthly 

intervals. Two of the replicated sets of ponds held only 

phytoplankton, whereas the third set of ponds contained 

rooted macrophytes. The authors found increased metabolism 

of penta in the ponds containing rooted macrophytes. The 

increased degradation of penta in the macrophyte ponds 

resulted in lower body burdens in channel catfish, bluegill, 

and largemouth bass. These fish species survived the highest 

dose of penta (1,000 μg/L) in the pond containing macrophytes but succumbed in the ponds containing only phytoplankton. The authors suggested three hypotheses as possible 

explanations for these results. However, only two of those 

hypotheses appear different from each other: (1) the presence 

of the macrophytes resulted in a different chemical or physical environment in the ponds that increased penta degradation, and (2) the macrophytes or aufwuchs community associated with the macrophytes were incorporating penta as a 

conjugate or within the cell structure. 

Fisher (1990) observed that the concentration of dissolved 

penta increased with increasing pH but that the uptake by 

both organisms and sediments decreased. She also observed 

that increasing sediment organic carbon was associated with 

higher sediment levels of penta and reduced concentrations 

in the water column. 

Delaune and others (1983) studied the fate of penta following 

a major spill in a Louisiana Gulf Coast estuary. They found 

that the degradation of penta was strongly influenced by 

sediment pH and redox potential. Degradation rates decreased with decreasing sediment redox and were maximized 

at pH 8.0, with reduced degradation at either higher or lower 

pH values. In addition, these authors found that penta was 

more tightly bound to oxidized sediment than to reduced 

sediment. They observed that all the penta had essentially 

disappeared from the spill area within 18 months and hypothesized that observed microbial degradation could account for the degradation. Sediments in the study area contained 3% to 5% carbon, and the sediment pH was essentially 

neutral at 6.8. 

It is apparent that a variety of microorganisms are able to 

degrade penta. In general, it appears that aerobic degradation 

is more efficient than anaerobic degradation and that increased degradation occurs at moderate temperatures (between 10o

C and 35oC). In most instances, it appears that 

penta degradation by microbial communities is enhanced by 

the presence of alternate carbon substrates. 

Metabolism 

Fisher (1990) noted that algae, invertebrates, and vertebrates 

rapidly metabolized penta. Biotransformation rates were 

significantly greater in algae, especially at higher pH levels 

and in association with high-TOC sediments. Biotransformation rates in snails and fish varied between 0.52 and 3.88, 

with no significantly different rate between the two phyla. 

Kukkonen and Oikari (1988) examined the metabolism of 

penta in the cladoceran Daphnia magna. The authors found 
45 

that neonate (<24 h old) and adult daphnia were equally able 

to metabolize penta and that humic content in the aquaria 

(dissolved organic carbon up to 23.5 mg/L) had no effect on 

the metabolic rate. The concentration of penta in these experiments was 20 μg penta/L, and the pH was low at 5.5. The 

authors did not determine a metabolic half-life. However, 

their data indicated that the concentration of free penta in the 

water column declined to 50% of the initial value in about 

10 h and that 50% of the penta taken up by Daphnia magna 

was metabolized in approximately 24 h. They concluded that 

the principle metabolic pathway in D. magna involved sulfate conjugation. 

Trujillo and others (1982) found the half-life of penta in 

midges to be 4.7 days. However, data in Lydy and others 

(1994) suggest that midge tissue concentrations of penta 

peaked at about 12 to 14 h, then declined rapidly to 50% of 

the maximum in about 24 h. The authors calculated a halflife of 15 h for penta in the midge. 

Glickman and others (1977) determined penta half-lives in a 

variety of rainbow trout (Oncorhynchus mykiss) tissues. 

The values varied from 6.2 h in blood to 23.7 h in fat. Penta 

is lipophilic, and these results suggest that there is short-term 

sequestration in body lipids. The authors found high penta 

levels in the bile of these fish and concluded that the penta 

was being conjugated with bile and excreted. 

Stehly and Hayton (1989) examined the metabolism of penta 

in rainbow trout, fathead minnows, sheepshead minnows, 

firemouth, and goldfish that had been exposed to penta for 

64 h. They found that penta metabolism was species specific. 

Consistent with other studies, Stehly and Hayton found that 

biliary excretion accounted for less than 30% of the total 

penta metabolites and that 76% of the penta metabolites in 

bile consisted of penta-sulfate or penta-glucuronide. All 

metabolites excreted into the water were sulfate conjugates, 

and bile was enriched in glucuronide conjugates. 

Similar results were demonstrated by Cravedi and others 

(1995) in Arctic char (Salvelinus alpinus) eleutheroembryos 

(end of yolk sac resorption or 50 to 100 mg wet weight). Test 

pH was 7.9 and duration was 48 h. They observed that pentaglucuronide accounted for 24.2% of the 14C found in water at 

the end of 48 h. The parent penta accounted for 29.5%, and 

pentachlorophenylsulfate represented 49.4% of the 14C present in the water column. 

Benner and Tjeerdema (1993) studied the toxicokinetics and 

biotransformation of penta in a marine species of fish (Atherinops affinis). The fish were exposed to 50 μg penta/L for 

24 h to determine the BCF, the elimination rate constant, and 

the elimination rate half-life. The absorption rate constant 

was 0.012/h, leading to a BCF of 278. The elimination rate 

constant was higher at 0.014/h, and an elimination half-life of 

52.7 h was determined. During 24 h of exposure to clean 

seawater, topsmelt depurated 32.9% of the retained penta and 

residues. Most (64.9%) of the penta was excreted unaltered. 

However, penta metabolites pentachlorophenylsulfate 

(18.9%) and pentachloro-β-D-glucuronide (16.2%) were also 

observed. The authors note that these same compounds have 

been identified as intermediates in the metabolism of penta 

by goldfish, fathead minnows, rainbow trout, firemouths 

(Cichlasoma meeki), sheepshead minnows, and striped bass. 

Benner and Tjeerdema concluded that topsmelt rapidly absorbed and more slowly depurated penta after short-term 

exposure through excretion and/or detoxification by sulfation 

and glucuronidation. 

Bioconcentration and Bioaccumulation 

Bioconcentration and bioaccumulation of contaminants are 

of special importance because some aquatic species, most 

notably bivalves, have demonstrated an ability to rapidly 

bioconcentrate contaminants in water to high tissue levels. 

The concern is that persistent contaminants may move up the 

food chain, biomagnifying to higher concentrations in each 

trophic level, until contaminants found at nontoxic levels in 

the ambient environment reach concentrations where they 

cause stress and disease. For effective biomagnification and 

movement of contaminants through the food chain, several 

conditions must be met. 

First, organisms must have the ability to bioconcentrate low 

levels of contaminants from the water column or their food. 

Second, these contaminants, or their toxic metabolic intermediates, must be retained, unaltered, in the tissues of the 

organism until it falls prey to an animal at a higher trophic 

level. 

Several factors mitigate biomagnification. If contaminants 

are not absorbed, they cannot accumulate. Numerous organisms, particularly vertebrates, have the ability either to metabolize or excrete organic contaminants. The gut, liver, 

kidney, and gall bladder are common sites of penta concentration, metabolism, and excretion in vertebrates. If the contaminants are either rapidly excreted or metabolized to nontoxic compounds, then the chain is broken and 

biomagnification is not effective in passing contaminants 

upward through the food chain. 

DDT is an excellent example of a persistent compound that 

was bioconcentrated from low levels in the water to higher 

levels—first in plankton, then in fish, and finally in bird 

populations, with devastating consequences. 

Bioconcentration of Penta From Water 

The uptake of penta is a function of pH (Fisher 1990, Fisher 

and Wadleigh 1986) and not so much of water concentration. 

For example, at pH 4.0 penta is fully protonated and therefore highly lipophilic, resulting in higher bioconcentration 

potential. Conversely, penta is completely ionized at pH 9.0 

with lower bioconcentration potential and significantly reduced toxicity. In general, Fisher (1990) observed a negative 
46 

correlation between the uptake of penta by algae, snails, and 

fish with pH. Highest bioconcentration factors (BCFs) were 

found at pH 4 (BCF = 117.2 to 681.9). Bioconcentration 

factors at environmentally realistic pH 6.0 and 8.0 ranged 

from 3.8 (algae at pH 8.0) to 271.1 (fish at pH 6.0). 

Maekelae and Oikari (1990, 1995) determined penta BCFs 

from145 to 342 in adult (55-mm valve length) freshwater 

mussels (Anadonta anatina). Tests were conducted at pH 6.5 

in penta concentrations of 7 and 14 μg/L. However, an equilibrium penta concentration of only 1.8 + 0.1 μg/L was 

reached in 4 and 16 h in the two experiments. They found 

higher penta concentrations in the digestive gland and kidney 

compared with whole body soft tissues. In a more recent 

experiment using (14C) penta, these same authors (Makela 

and Oikari 1995) found steady state BCFs averaging 100 in 

Anadonta anatina and 73 in Pseudanodonta complanata. 

The pH in these experiments was 6.5, and the penta concentration was 9.7 μg/L. The BCFs were determined as the body 

burden of penta (measured in a number of ways) divided by 

the final concentration of penta. It must be recognized that 

when determining BCFs for labile chemicals, such as penta, 

the average exposure affecting uptake is likely much greater 

than the final concentration, which tends to inflate the BCF. 

It would appear that the development of biologically meaningful BCFs should use some intermediate concentration of 

the contaminant. An appropriate protocol would necessarily 

require consideration of the dynamics of degradation in the 

ambient water integrated over time compared with the depuration and metabolism of the contaminant in the organism at 

question. 

Niimi and Cho (1983) suggested that penta is rapidly accumulated and eliminated by trout in the natural environment. 

Uptake from water appeared to be the most important pathway, and accumulation through food was thought to represent 

only a minor contribution. Consistent with Glickman and 

others (1977) and Rogers and others (1990), Niimi and Cho 

observed highest penta levels in the liver and bile of fish fed 

penta-contaminated feed to satiation for 40 days. Biliary 

excretion appeared to be a major route of depuration. They 

noted that the residence time of penta in water was short and 

suggested that its impact on fish would be most evident in 

localized areas that receive a continuous input of penta from 

a point source. Furthermore, Niimi and Cho (1983) concluded that bioaccumulation (biomagnification) of penta 

through the food chain is minimal and the observation that 

penta levels in smelt and alewive, the primary forage species 

for many Lake Ontario salmonids, were similar to concentrations found in the predators. 

Bioaccumulation of Penta From Sediments 

The ultimate fate of penta deposited in aquatic environments 

is either decomposition in the water column or sedimentation. Fisher (1990) concluded that “Thus, for the organic 

sediment system, bioaccumulation will be determined by 

interactions between pH, available sorption sites, degree of 

ionization of PCP, and levels of sediment ingestion.” 

Midges are deposit feeders and are known to rework sediments by feeding and burrowing. Therefore, midges are not 

only exposed to penta in interstitial water, but they also 

ingest contaminated particles. Fry and Fisher (1990) compared the bioaccumulation of penta by Chironomus riparius

allowed to burrow in penta-contaminated sediments with 

similar midges held in suspension directly above the sediments. A third experiment exposed dead midge larvae to 

contaminated sediments to evaluate the passive uptake of 

penta. The authors documented BCFs of 229 from water and 

7.3 from sediment in which the midges were burrowing. 

The bodies of dead midges exposed to contaminated sediment yielded a BCF of 13.3. Fry and Fisher (1990) noted that 

Trujillo and others (1982) found the half-life of penta in 

midges to be 4.7 days. They hypothesized that the lack of 

metabolic degradation in the passive uptake experiment 

contributed to the higher ultimate penta body burden in the 

dead chironomids. However, they noted that ionized penta 

appeared to have a significant affinity for the body wall of 

the dead larvae and concluded that passive uptake from pore 

water appeared to be important to the accumulation of penta 

from sediments. 

Fry and Fisher (1990) noted that penta has a log Kow of 5.01. 

They concluded that penta does not behave like a neutral 

lipophilic compound and therefore its activity and fate are 

not predictable from its octanol–water partition coefficient. 

In contrast, Lydy and others (1994), found that BCFs in the 

midge Chironomus riparius (BCF = 458) were reasonably 

well predicted by a much lower octanol–water partition 

coefficient (Kow = 0.758) when the water concentration was 

9 μg penta/L. 

Fisher (1990) found that the sorption of penta to organic 

sediment significantly reduced its concentration in the water 

column at all pH levels, thereby reducing accumulation of 

penta in the microcosm’s organisms. At environmentally 

realistic pH 6 to 8, BCFs were lower in microcosms having 

3% TOC sediments compared with microcosms having sediments lacking any TOC. 

Haque and Ebing (1988) examined the bioconcentration of 

penta from water and soil. They determined BCFs from soil 

of 6.3 for Allolobophora caliginosa and 22.2 for Lumbricus 

terrestris. They found that penta was rapidly bound to the 

soil and concluded that ingestion of contaminated particles 

was a significant pathway. 

Van Gestel and Ma (1988) investigated the toxicity and 

bioaccumulation of penta in the earthworms Eisenia fetida 

andrei and Lumbricus rubellus in organically rich sandy soils 

(3.7% to 6.1% organic matter) with low pH (5.0 to 5.6). 

The LC50 values for Lumbricus rubellus were 883 mg 
47 

penta/kg dry soil in Holten soil (pH 5.6 and 6.1% organic 

matter) and 1,094 mg penta/kg dry soil in Kooyenburg soil 

(pH ~5.0 and 3.7% organic matter). The differences in these 

LC50 values were not significant. Bioconcentration factors 

were based on the average of the sediment values on day 0 

and day 14. This appears a more reasonable approach than 

using the values on the last day, which was done by Maekelae and Oikari (1990). The BCF, based on bulk sediment 

penta concentration, varied between 3.4 and 8.0. In contrast, 

the BCFs varied between 426 and 996 when based on porewater penta concentrations. The authors suggest that the 

porewater BCF was consistent with BCFs of 475 observed in 

fish. 

Biomagnification of Penta in Food Chains 

Schuytema and others (1993) fed mealworms contaminated 

with 64.8 to 2,604 μg penta/g to African clawed frogs 

(Xenopus laevis) for 27 days. They observed no mortality in 

the frogs and no significant bioaccumulation of penta. Highest concentrations of penta were found in the frog’s liver. 

However, these levels were inversely proportional to the 

level of penta in the mealworms. Penta was found in frog 

liver at <4.6 μg/g in frogs fed mealworms contaminated to 

64.8 μg penta/g, whereas the liver of frogs fed mealworms 

with 2,604.6 μg penta/g contained penta at <0.6 μg/g. 

Niimi and Cho (1983) determined the uptake from food and 

half-life of penta in rainbow trout (Oncorhynchus mykiss). 

Trout fed diets containing 40 μg penta/kg food attained 

whole body levels of 2 μg/kg during the 3-month study. 

Trout fed with penta-contaminated feeds with levels 75 times 

greater (3,000 μg/kg) accumulated 40 μg penta/kg at the end 

of 40 days. This level declined to 20 μg/kg by the end of the 

study. The biological half-life of penta in trout was estimated 

to be approximately 7 days. This observation is consistent 

with the literature describing the depuration and metabolism 

of penta in vertebrates and invertebrates. It appears that penta 

clearance in all tested organisms is fast enough to minimize 

any potential for the compounds to biomagnify in food 

chains. 

Toxicity 

Penta is known to uncouple oxidative phosphorylation, inhibiting adenosine triphosphate (ATP) pathways important to 

respiration in both animal and plant cells. In addition, Moreland and Hilton (1976) described penta as a general inhibitory uncoupler, suggesting that it has several sites of action, 

including photophosphorylation, protein synthesis, and lipid 

biosynthesis (Morrod 1976). All the mechanisms of penta’s 

toxicity have not been precisely defined but may generally 

involve the disruption of cellular membranes (Jayaweera and 

others 1982, Senger and Ruhl 1980, Smejtek and others

1983). 

Acute toxicity causes observable physiological lesions and is 

usually measured by mortality. Penta interferes with the 

oxidative phosphorylation by uncoupling the production of 

adenosine triphosphate from adenosine diphosphate. Because 

this process provides the energy source for cellular metabolism in most organisms, penta is a broad-spectrum biocide. 

A common measure of acute toxicity is the concentration of a 

toxicant that causes 50% mortality in a test population within 

some specified period (often 96 h). This parameter is referred 

to as the 96-h LC50. Eisler (1989) summarized 96-h LC50 

concentrations for aquatic organisms. For most freshwater 

species, the 96-h LC50 varied between 100 and 2,000 μg 

penta/L. In general, Eisler’s (1989) data suggest that 

freshwater vertebrates (fish) are more sensitive than 

invertebrates. Table 21 summarizes the lower LC50 values 

provided by Eisler (1989) for salmonids and centrarchids. 

Salmonids of the genus Oncorhynchus appear most sensitive. 

In contrast, invertebrate LC50 values are typically greater 

than 100 μg/L. 

Chronic Toxicity 

Toxicants can have subtle effects that are important to the 

competitiveness of individuals in natural environments and 

the sustainability of populations of organisms. Numerous 

endpoints are evaluated in assessing chronic effects. Most 

commonly, these endpoints involve reproduction and/or 

growth. The lowest contaminant concentration in a bioassay 

that does not produce an observable effect is referred to as 

the no observed effect concentration (NOEC). The lowest 

concentration at which the effect being evaluated is observed 

is referred to as the lowest observed effect concentration 

(LOEC). If an effect is observed at 200 μg/L but not at 

100 μg/L, the first value would be reported as the LOEC and 

the second as the NOEC. The actual effects threshold would 

be somewhere between the two values. ENVIRON (1996) 

summarized a number of chronic endpoints for aquatic fauna 

and flora exposed to penta. In Table 22, data are further 

summarized to include only those that included the pH test. 

Table 21—Acute toxicity of freshwater fish to penta 

(96-h LC50)

a

 

Species 

Concentration 

(μg penta/L) 

Oncorhynchus mykiss (rainbow trout) 34 to 121 

Oncorhynchus nerka (sockeye salmon) 63 to 68 

Oncorhynchus tshawytscha (chinook 

salmon) 

68 to 78 

Salmo salar (Atlantic salmon) 500 

Salvelinus fontinalis (brook trout) 128 

Lepomis macrochirus (bluegill) 120 to 350 

Micropterus salmoides (largemouth bass) 136 to 287 

a

Reported in Eisler (1989). 
48 

Roszell and Anderson (1994) examined the effect of penta on 

nonspecific immune function in two phagocytic cell populations isolated from the estuarine fish, Fundulus heteroclitus. 

They found that phagocytosis of yeast particles was significantly inhibited at penta concentrations greater than 

5,000 μg/L. A comparison of phagocytic response between 

controls and a penta level of 1,000 μg/L did not reveal significant differences. 

Brown and others (1987) observed a reduced number of 

feeding acts in the young of largemouth bass exposed to 

penta concentrations of 67 and 88 μg/L but not at concentrations less than 67 μg/L. Endpoints measured included the 

number of feeding attempts and the number of misses and 

mistakes. The NOEL was 45 μg penta/L. Hardness (65 mg/L 

calcium carbonate) and pH (7.7) were not measured directly 

in this experiment but were assumed equal to that observed in 

a 1985 study using the same water supply. 

Keller (1993) noted that in 1993 more than 40 species of 

freshwater unionid mussels were listed as endangered or 

threatened under the Endangered Species Act. Keller determined a 48-h LC50 value in juvenile Anadonta imbecilis at 

pH 7.0 and compared the results with concurrent bioassays 

on Daphnia magna and Lepomis macrochirus. Results indicated that the 48-h LC50 for the juvenile mussels (610 μg 

penta/L) was greater than the 48-h LC50 for the daphnid 

(330 μg/L) or the 96-h LC50 for bluegills (240 μg/L). At pH 

7.0, the EPA chronic penta water quality standard is 5.73, 

giving a safety factor of 106. 

Temperature Effects 

Fisher and Wadleigh (1986) examined the toxicity of penta 

to the midge Chironomus riparius at 15o

C, 25oC, and 35oC. 

The endpoint she examined was a flight response following 

stimulation with a pair of forceps during exposure to PCP in 

soft water at pH 7.0. Fisher observed that EC50 values increased from 1,176 μg/L at 15o

C to 1,556 at 25oC, then 

declined to 631 μg PCP/L at 35o

C. She concluded that midge 

metabolism increased at the higher temperatures and that 

penta’s interference with respiration (phosphorylation of 

Table 22—No observed effects level (NOEL) and lowest observed effects level (LOEL) associated with penta in 

freshwater environmentsa 

Endpoint Duration Species NOEL LOEL pH °C EPA 

Reproduction; number viable eggs 16 Lymnaea stagnalis (snail) 50 NR 8.0 18.3 46.5 

Larval survival and reproduction 10 – 28 American Flagfish 55 102 6.95 25 5.4 

Biomass & mortality, eggs @ 10o

C; 

alevins @ 15 o

C; fry @ 20 oC 

>28 Oncorhynchus mykiss 10.9 25 8.0 10 – 20 15.6 

Hatchability, survival, & growth 32 Pimephales promelas 16.5 34.6 6.5 25 3.5 

Survival & growth (fry & juveniles) 90 Pimephales promelas 6 13 7.4 25 8.6 

Survival & growth (fry & juveniles) 90 Pimephales promelas 36 85 7.4 25 8.6 

Survival & growth (fry & juveniles) 90 Pimephales promelas >130 >130 9.4 25 63.9 

Early life stage hatchability, survival, & growth 32 Pimephales promelas 44.9 73 7.55 25 10.0 

Hatchability, survival, & growth 32 Pimephales promelas 63.7 125 8.5 25 25.9 

Hatchability, survival, & growth 32 Pimephales promelas 27.6 58.2 7.5 9.5 

Hatchability, surviva,l & growth 32 Pimephales promelas 32 75 8.0 15.6 

Growth 56 Chaetogammarus marinus 100 NR 8.0 NR 15.6 

Number of viable oocytes 18 Oncorhynchus mykiss 11 19 7.4 12 8.6 

Number of viable oocytes 18 Oncorhynchus mykiss 12 22 7.5 12.5 9.5 

Reproduction 21 Daphnia magna 180 320 8.0 20 15.6 

Survival and reproduction 7 Daphnia magna 100 500 8 20 15.6 

Inhibition of cell growth 5 Skeletonema costatum 11 20 8.1 19 – 22 17.3 

Inhibition of cell growth 5 Selenastrum capricornutum 12 17 7.5 24 – 25 9.5 

Inhibited cell growth 5 Anabaena flos-aquae 7.8 18 7.5 24 – 25 9.5 

Growth, reduction; biomass 21 Elodea canadensis 230 380 7.95 22 14.9 

Inhibited cell growth 5 Inavicula pelliculosa 40 77 7.5 25 9.5 

Inhibited cell growth 5 Anabaena flos-aquae 7.8 18 7.50 24 – 25 9.5 

Frond density & biomass 14 Lemna gibba 32 72 5.0 23 – 27 0.8 

a

All penta values are in μg penta/L. 
49 

ADP to ATP) at the higher metabolic rate was responsible 

for the increased EC50 at 35o

C. Similar results were reported 

by Fisher (1991) in an experiment conducted at pH 4, 6, and 

8. The EC50 values for penta in this study varied between 253 

μg/L at pH 4 and 35o

C to 2,052 μg penta/L at pH 8 and 25oC. 

Eisler (1989) cites similar results from Hedtke and Arthur 

(1985) who reported positive correlations between EC50 and 

temperature for fathead minnows (Pimephales promelas), the 

isopod Asellus racovitzai, and the snail Physa gyrina. 

In summary, it appears that the toxicity of penta to aquatic 

species increases with increasing temperature. Similar increases in the toxicity of penta to Notopterus notopterus with 

increasing temperature (16 oC, 23 oC and 36o

C) were observed by Gupta and others (1983). 

pH Effects 

Fisher (1991) demonstrated that the toxicity of penta is inversely related to pH. The effective concentration (EC50) of 

penta resulting in failure of Chironomus riparius to execute 

an appropriate flight response increased from 384 μg/L at 

25o

C and pH 4 to 2,052 μg penta/L at the same temperature 

and pH 8.0. 

Smith and others (1987) examined the toxicity of penta to 

Selenastrum capricornutum and found that culture media 

equilibrium pH and 96-h EC50 were strongly correlated 

(r = 1.00) between pH 7.3 and 8.5. The authors concluded 

that the toxicity of penta is due primarily to the concentration 

of the undissociated compound. 

The dependence of penta toxicity on pH was further elucidated by Spehar and others (1985) in fish and amphipods. 

They found that acute exposures in all three species showed 

that penta toxicity was decreased with increasing pH. Kaila 

and Saarikoski (1977) observed a similar response in crayfish 

(Astacus fluviatilis). The 8-day LC50 decreased from 53 mg/L 

at pH 7.5 to 9 mg penta/L at pH 6.5. 

Stehly and Hayton (1990) described the uptake and clearance 

of penta in goldfish (Carassius auratus) as a function of 

environmental pH (7.0, 8.0, and 9.0). The authors found 

reduced uptake and clearance at increasing pH. They concluded that pH-related changes in the pharmacokinetics of 

penta resulted in a decrease in its BCF with increasing pH 

and suggested that this could account for both the decreased 

capacity of fish to accumulate penta and its reduced toxicity 

at higher pH values. 

Early life-stage exposures of fathead minnows showed that 

chronic penta toxicity and bioaccumulation were similarly 

decreased when pH values were increased. Stehly and Hayton developed a relationship describing bioconcentration as a 

function of pH. Reported bioconcentration values ranged 

from 1,066 at pH 6.5 to 281 at pH 8.5: 

Fathead minnow BCF = 10[4.80 – 0.28(pH)] (R2

 = 0.94) 

The authors concluded that the decrease in chronic penta 

toxicity appeared to be due to reduced bioaccumulation and 

toxicity as a direct result of the increased dissociation at 

higher pH values. 

Water Hardness Effects 

Inglis and Davis (1972) examined the effects of water hardness (13.0, 52.2, 208.7, and 365.2 mg calcium carbonate/L) 

on six species of fish, including rainbow trout. Reported 

values of pH ranged from 7.8 at a hardness of 13.0 mg/L to 

pH 8.0 at all other values of hardness. The authors concluded 

that water hardness had no significant effect on the toxicity 

of penta to any of the tested species. 

Dissolved Organic Carbon Effects 

Lee and others (1993) examined the acute toxicity of penta to 

zebrafish (Brachydanio rerio) and the cladoceran (Daphnia 

magna) at TOC concentrations varying between 0.0 and 50 

mg/L. They observed no significant difference in the 96-h 

EC50 (zebrafish bioassay) or 48-h EC50 (Daphnia magna) as 

a function of TOC at any of the tested values. 

Toxicity to Aquatic Plants 

Smith and others (1987) investigated the toxicity of penta to 

Selenastrum capricornutum as a function of pH and found 

that the 96-h EC50 was given by the relationship 

96-h EC50 (Selenastrum capricornutum) 

 = exp[0.847(pH) − 4.28] 

The regression coefficients in this relationship are similar to 

the current EPA penta standard and demonstrate reduced 

toxicity at elevated levels of pH where more penta is in the 

dissociated form (that is, having lost the OH radical). The 

authors suggested that the toxicity of penta is primarily associated with the undissociated species of the compound. 

Regulatory Criteria for the Protection 

of Aquatic Resources 

Section 304 of the Clean Water Act requires the EPA to 

publish and update ambient water quality criteria (USEPA 

1980a). These criteria reflect the latest scientific knowledge 

of the identifiable effects on the health and welfare of aquatic 

resources including fauna, flora, and human uses. 

The brief review provided in this document is consistent with 

U.S. EPA (USEPA 1986) in predicting that penta acute and 

chronic toxicity are inversely correlated with water pH and 

dissolved oxygen concentration and directly correlated with 

temperature. At pH 6.5, the U.S. EPA found that acute values 

ranged from 4.4 to greater than 43,920 μg penta/L. Chronic 

values ranged from less than 1.835 to 79.66 μg penta/L. The 

mean acute–chronic ratios ranged from 0.89 to 15.79 μg/L. 

Freshwater algae were affected by concentrations as low as 

7.5 μg/L, whereas vascular plants were adversely affected at 
50 

greater than 89 μg/L. Bioconcentration factors ranged from 

7.3 to 1,066 in three species of fish. 

Procedures described in the Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses (USEPA 1985) indicate 

that except where a locally important species is highly sensitive to penta, freshwater aquatic organisms and their uses 

should not be affected unacceptably if the average short-term 

concentrations (μg penta/L) do not exceed the numerical 

values given by 

Acute criterion:

 1-h average concentration 

 < exp[1.005(pH) – 4.830] μg/L 

Chronic criterion:

 4-day average (once every 3 years) 

 < exp[1.005(pH) – 5.290] (μg/L) 

Note that these are not appropriate concentrations for continuous exposure, and they are not no-observed-effect levels 

or threshold effects levels. For instance, at pH 6.8, the 

chronic criterion for penta is 4.68 μg/L. At this pH, the U.S. 

EPA (USEPA 1986) notes that a penta concentration of 

1.74 μg penta/L caused a 50% reduction in growth of yearling sockeye salmon in a 56-day test. This may seem inconsistent with the chronic EPA value. However, remember that 

the EPA criterion is for a maximum 4-day exposure, not a 

56-day exposure. 

These criteria are not rules, and they do not have regulatory 

impact. However, several states have used the U.S. EPA 

(USEPA 1986) as the basis for setting regulatory standards 

for potentially toxic compounds, including penta: 

Washington State (WAC 173-201), Texas (30 TAC 307.6), 

Utah (UAC R317-2-14), Florida (FAC 62-302.530), Indiana 

(327 IAC 2-1-6), and Oregon (OAR 340 41). 

Different criteria have been proposed or adopted by other 

jurisdictions. The most prominent of these were reviewed by 

ENVIRON (1996). Data for continuously distributed criteria 

provided in ENVIRON (1996) were subjected to nonlinear 

regression analyses. The results are presented in Table 23 

and Figure 23. 

The Draft 1995 CCME criteria and the 1995 Ontario guidelines (ENVIRON 1996) do not adequately reflect the effect 

of pH on the toxicity of penta. These criteria are either under 

protective at low pH or over protective at high pH. The 1994 

Aquatic Risk Assessment and Mitigation Dialogue Group 

(ADG) Level of Concern Approach (ENVIRON 1996) produces two sets of criteria. The lower values are sufficient to 

protect 99% of the species, and the upper values are sufficient to protect 95% of the species. These are continuous 

criteria; therefore, it is inappropriate to compare their recommendations with those of the 1-h and 96-h U.S. EPA 

criteria. 

In response to recommendations made by an EPA task force, 

the ADG was sponsored by the U.S. EPA in 1992 and the 

North American Chemicals Association in 1993. The ADG 

included representatives of the U.S. EPA, agrochemical 

companies, academia, and environmental and agricultural 

interest groups. The Society of Environmental Toxicology 

and Chemistry (SETAC) acted as a facilitator for meetings of 

the group and published a final report with recommendations 

(SETAC 1994). The ADG recommended an “integrated 

probabilistic risk assessment approach that included both the 

Table 23—Various jurisdictional criteria for pentaa 

Jurisdiction Criterion Basis for criterion 

Proposed 1995 interim 

CCME (1997) guidelines 

0.02 at pH 6.5; 0.1 at pH = 7.0 or 7.5; 

0.30 at pH > 8.0 

Extrapolation from acute guppy data; 

safety factor = 0.1 

ENVIRON (1996) calculated Chronic criterion = exp[1.002(pH) – 7.90] Application of 0.1 safety factor to lowest values 

using 1991 CCME chronic LOEC in the database 

Ontario Canada 

 ENVIRON (1996) 

0.5 at all values of pH Application factor of 0.01 applied to the lowest 

mean species-specific acute LC50 independent of 

pH 

U.S. EPA (1986) 4-day chronic = exp[1.004(pH) – 5.28] Statistical approach using all relevant acute toxicity data; set to protect 95% of species; ACR = 

3.166 

Great Lakes Water Quality 

Initiative (ENVIRON 1996) 

4-day chronic = exp[1.004(pH) – 5.12] Similar to U.S. EPA approach except the 

ACR = 2.608 

SETAC (1994) Continuous chronic = exp[1.005(pH) – 7.26] Graphical approach using entire database of 

Group Level of Concern primary chronic NOECs 

a

The U.S. EPA and Great Lakes chronic criteria are for 4-day exposures. Others listed are continuous criteria. 

 All values are in μg/L. 
51 

Figure 23—Various jurisdictional criteria defining allowable concentrations of penta in 

freshwater as a function of average pH. All values are in micrograms penta per liter. 

probability of exposure and effects.” ENVIRON (1996) 

applied the ADG methodology to derive penta chronic toxicity guidelines sufficient to protect 90%, 95%, and 99% of 

species. The results for the protection of 95% and 99% of 

species are included in Table 23 and Figure 23. 

Carcinogenicity 

Jorens and Schepens (1993) note that penta is not classified 

as a human carcinogen but suggest that historical contaminants in commercial penta products, such as chlorodibenzop-dioxins, are known carcinogens. The EPA has classified 

penta as a B2 carcinogen (that is, no evidence of carcinogenic response in humans, but evidence in animals is sufficient to cause the compound to be suspect). Similarly, Health 

Canada (1994) and IARC (1991) have classified penta as a 

Group III.B compound that is possibly carcinogenic to 

humans. 

Mutagenic and Teratogenic Effects 

Venegas and others (1993) examined the teratogenic effects 

of penta in an amphibian (Caudiverbera caudiverbera). They 

examined embryonic development in water containing 15, 

150, 300, and 1,500 μg/L penta. Results of the micronucleus 

test on premetamorphic larvae indicated higher micronucleus 

formation rates in the controls than in the various penta 

treatments. Chromosomal aberration tests in anaphase and 

telophase provided a second endpoint. The authors concluded that the highest penta dose resulted in inhibition and 

delays in development of normal growth of embryos of this 

amphibian. However, the study found no clear mutagenic 

effects. CCME (1997) concluded that although chlorophenols may have reproductive and fetotoxic effects, they do not 

appear to be teratogenic or mutagenic. 

Toxicity Summary 

For the purpose of this risk assessment, we used the U.S. 

EPA chronic criteria as a benchmark against which to compare levels of penta migrating into the water column during 

the first 4 days. The 1994 ADG, necessary to protect 99% of 

species, was used as a second benchmark against which to 

assess long-term water column concentrations of penta within 

a few centimeters of the immersed wood on days >5.0 following immersion of the treated wood. 

Penta Sediment Quality Benchmarks 

Washington State (Washington Administrative Code, Chapter 173-204-320) has developed an apparent effects threshold 

(AET) penta standard for marine sediments at 360 μg/kg. 

The AET is the lowest concentration of a chemical above 

which adverse effects are always observed in Puget Sound 

sediments. Adverse effects are determined from laboratory 

bioassays on a variety of test animals and on the paired 

analysis of infaunal communities and sediment chemistry. 

These standards are considered sufficient to protect most 

marine organisms. 

10

20

30

40

50

60

pHPenta criteria (

μg/L)

1995 CCME/BC 0.02 0.1 0.1 0.3 0.3 0.3 

ENVIRON 0.25 0.41 0.68 1.12 1.85 3.06 

Ontario 1995 0.50 0.50 0.50 0.50 0.50 0.50 

EPA Chronic 3.47 5.73 9.47 15.65 25.86 42.71 

Great Lakes 4.06 6.70 11.07 18.29 30.21 49.90 

1994 ADG (99%) 0.48 0.80 1.32 2.18 3.61 5.96 

1994 ADG (95%) 1.33 2.20 3.64 6.01 9.93 16.40 

6.5 7 7.5 8 8.5 9 
52 

The New York State Department of Environmental Conservation (NYSDEC 1993) has established a freshwater sediment criterion for penta of 100 μg penta/g sedimented organic carbon for acute toxicity and 40 μg penta/g sedimented 

organic carbon to prevent chronic toxicity. These criteria are 

based on an equilibrium partitioning model and state water 

quality criteria for penta and are intended for use in screening 

contaminated sediments. The NYSDEC criteria are expressed as a concentration in sediment by prescribing an 

organic carbon content. For example, in a sediment with 1% 

organic carbon, the corresponding chronic criterion value is 

400 μg penta/kg dry sediment (40 μg/g TOC × 0.01 

TOC = 0.4 μg/kg), or 800 μg/kg at 2% TOC, 1,200 μg/kg at 

3% TOC, etc. 

The Washington State marine standard of 360 μg penta/kg 

dry sediment weight was used as a benchmark in assessing 

marine environmental risks associated with the use of pentatreated wood, and the NYSDEC criteria was used for freshwater sediments. The risk assessment program automatically 

computed the sediment penta criteria based on a user input 

for TOC. 

Loss of Penta 

Southern yellow pine piling was treated to 8.0 kg/m3

(0.5 lb/ft3

) of penta (in the treated zone) and immersed in 

freshwater and salt water at a variety of pH values. Penta loss 

rates used in this assessment were developed data from a 

1994 study, Pentachlorophenol – Leaching From Utility 

Poles Exposed to the Aquatic Environment, prepared by 

Springborn Laboratories for the Pentachlorophenol Task 

Force as part of data call-in for re-registration of penta by the 

U.S. EPA (Connor 1994). The study was produced and 

compiled in accordance with all pertinent U.S. EPA Good 

Laboratory Practice regulations. 

The utility poles chosen for testing were selected from a kilndried stock of southern yellow pine. Two utility pole sections, approximately 3.3 m (11 ft) long and 25.4 cm (10 in.) 

in diameter were treated to 8.0 kg/m3

 (0.5 lb/ft3) of penta in 

the treated zone using the Lowry Process. An additional 

3.3-m (11-ft) untreated utility pole section was selected from 

the same kiln lot as a control. The three poles were cut into 

30 segments. 

The pole segments were leached in tanks containing unbuffered reagent water, reagent water buffered at pH 5, 7, 

and 9, seawater, and 0.1 N hydrogen chloride solutions. 

Triplicate tanks of each solution contained a treated pole 

segment, with an additional tank for each solution containing 

an untreated control. In each tank, the pH was measured 

daily and adjustment made with either acid or base to maintain the pH within +0.1 units. Leachate temperature varied 

between 18.3o

C and 22.0oC during the experiment, with a 

mean and standard deviation of 20 + 0.2o

C. The air-exposed 

cut end of the pole sections was sealed with epoxy, and 

TENAX traps (TENAX SpA, Italy) were installed in the 

sealed tanks to capture any volatalized penta. The traps did 

not reveal evidence of the volatilization of penta during the 

study. 

The immersed, cut ends of the poles were not sealed. Haloui 

and others (1995) determined the diffusivity of penta along 

three orthogonal axes in treated wood. They found that the 

longitudinal diffusivity (parallel to the grain structure) was 

seven times greater than either the radial or tangential diffusivity. This has implications for leaching studies, because if 

the end of the pole is exposed to water, as it was in this case, 

it can represent a significant fraction of the total leaching 

surface area. Penta released from the cut end of a pole or 

piling that is driven into sediments about 3 m will not be 

available to the water column. Therefore, in estimating the 

preservative lost to the water, only the surface grain should 

be considered. A pole diameter of 25.4 cm (10 in.) would 

result in a cut-end surface of 506.7 cm2

. If the length of the 

pole was 27.8 cm (11 in.), then the horizontal grain surface 

area exposed to the water would be 2,218.2 cm2

. If preservative is lost seven times faster at the cut end, then the loss 

rates determined in this study should be reduced to 47% of 

the observed rate [(506.7 + 2218.2) Observed Rate of Loss ÷ 

(7 × 506.7 + 2218.2) = Longitudinal Surface Loss Rate = 

0.47 × Observed Loss Rate]. 

There is a second problem associated with conducting static 

leaching tests of highly degradable organic compounds such 

as penta. The study did not state that the tanks were held in 

the dark. However, no information was given describing the 

lighting conditions (intensity or wavelength) under which the 

glass aquariums were held. The literature describing the 

degradation of penta clearly demonstrates that in normal 

sunlight, penta half-lives can be several days, certainly less 

than the 30 days over which this experiment was conducted. 

Therefore, some photolysis of the penta should be anticipated 

during the experiment. Secondly, both fungi and bacteria 

have demonstrated an ability to degrade penta (McAllister 

and others 1996). Connor (1994) noted the presence of fungi 

in the pH 5 and 7 treatments after approximately 1 week of 

immersion. Bacterial growths were reported in all tanks in 

this study. It is likely that these growths were metabolizing 

penta following the first week or two of immersion. 

The antagonistic effects associated with increased loss rates 

from exposed end grain and decreased loss rates associated 

with the metabolism of penta by microbes and degradation 

by light were not quantifiable in this experiment, and no 

correction was attempted. However, note that the increases as 

a result of not sealing the end grain would be observed immediately as would decreases associated with photodegradation. The decreases in calculated loss rate associated with 

microbial catabolism would not be observed for probably 

5 to 10 days. As shown in the following paragraphs, this risk 

assessment focuses on the first 5 days of immersion because 

that is the period in which maximum penta concentrations are 

anticipated in the water column. Therefore, it appears that the 
53 

assessment may be somewhat conservative in that greater 

loss rates associated with the exposed end grain were not 

being balanced by microbial degradation in the near term. 

The leachate was analyzed following 1, 3, 7, 14, 21, and 

30 days of immersion using high performance liquid chromatography (HPLC). As previously noted, the study followed 

EPA Good Laboratory protocols. Note that between 99.9% 

and 108% of the penta added to spiked samples was recovered in this study. 

Data provided by Connor (1994) were analyzed in the Statistica Non-Linear Estimation algorithm and results were used 

to develop loss rates per square centimeter per day. The 

sampling times were converted to the mean day during the 

sample period rather than the day on which samples were 

collected. The resulting predictive equation is provided. The 

regression explained 73% of the variation in the database, 

and each of the final coefficients were significant at α = 0.05. 

Penta loss = 10.9 exp[−0.255(day) + 0.355(pH) 

 + 0.01(salinity)] μg/cm2

-day (6) 

This expression was used to estimate the loss of penta from 

poles treated to 8.0 kg/m3

 (0.5 lb/ft3) in the treated zone. 

Dioxins in Penta 

Penta does not contain 2,3,7,8 tetrachloro-dibenzo-dioxin 

(2,3,7,8 TCDD). The U.S. EPA has established water quality 

criteria for this compound of <0.01 μg/L maximum and 

0.00001 μg/L continuous concentrations. 

Penta does contain less toxic hepta- and hexa-dioxins 

(HxCDD) for which no water quality standards were 

available. It is possible, using the toxic equivalency factors 

(with respect to TCDD) for these compounds and summing 

over their concentration in commercial penta to express the 

sum in terms of a toxic equivalency quotient (TEQ). The 

toxic equivalency for penta is 2.84 mg/L. This risk assessment assumes that the hepta- and hexa-dioxins found in penta 

are lost to aquatic environments in proportion to their relative 

proportion in the preservative (2.84 mg dioxin/L of preservative). The resulting value can then be compared with the EPA 

standards for 2,3,7,8 TCDD in assessing risk. 

Results From Penta-Treated Cougar Smith Bridge 

A total of 11 penta-treated bridges were considered as candidates for this risk assessment. All but four of these were 

eliminated because they also contained creosote-treated 

components. The Cougar Smith Bridge (Washington state) 

crossing the West Fork of the Satsop River is an HS20 structure with a clear span of 60.96 m (200 ft). This bridge was 

constructed in 1996. All wood components of this bridge 

were preserved to a retention of 8.0 kg/m3

 (0.5 lb/ft3) (in the 

treated zone), with penta dissolved in light solvent oil. The 

bridge sits on concrete footings, and the treated wood is not 

immersed in water. The Satsop River is used extensively by 

several species of salmon. Salmon were observed spawning 

in the gravel under and just downstream from the bridge in 

the fall of 1997. 

The Cougar Smith Bridge was originally scheduled for survey in the fall of 1997. However, an early series of storms 

raised the Satsop River above flood stage, and the monitoring 

was rescheduled for August 4, 1998, to allow for possible 

preservative loss associated with migration of the solvent 

oil during summer heating. The location of the bridge is 

Figure 24—Location of the Cougar Smith Bridge crossing the West Fork Satsop River in 

Southwestern Washington. 

 Cougar Smith Bridge 
54 

described in Figure 24. Figure 25a depicts the entire structure, and Figure 25b is included to illustrate the large quantity of treated wood involved in the bridge decking and support beams. 

As shown in Figure 26, the Satsop River was experiencing 

moderate to low flows during the August 4, 1998, survey. An 

area of the river with slowest flows and relatively sandy 

sediments was chosen for the sampling transect because this 

area appeared most likely to contain elevated levels of the 

preservative. Current speeds averaged 27 cm/s along the 

sampling transect. Currents in the main channel averaged 

63.7 cm/s in water averaging a 36-cm depth. 

During this survey, the weather was clear and sunny with no 

wind. The water temperature was 13.5o

C, and the air temperature was 23.9o

C at 1100 hours. Water pH was essentially 

neutral, averaging 6.90 + 0.1 (three measurements). Water 

column alkalinity was 33.25 and hardness was 28.16. Conventional physicochemical characteristics of sampled sediments are described in Table 24. Sediments along the sampled transect were dominated by gravel and cobble with 

minor amounts of sand, silt, and clay. Sediment total 

volatile solids (TVS) averaged 1.74 + 0.16%, equivalent to 

1.04% TOC. 

Sediment and Water Column Concentrations 

Penta was not detected in the water column at the upstream 

control station or under the bridge at the detection limit of 

0.25 μg/L (three samples each station) (Table 24). Penta was 

detected in one sediment sample located 0.45 m (1.5 ft) 

downstream from the downstream perimeter of the bridge at 

a level of 19.0 μg/kg. The sediment concentration at the 

1.8 m (6.0 ft) downstream station was estimated at 6.6 μg/kg, 

which was less than the calculated detection limit. Penta was 

not detected in other samples at this bridge. 

Sediment Benchmark 

As previously noted, the New York State Department of 

Environmental Conservation (NYSDEC 1993) has established a freshwater sediment criterion for penta at 

40 μg penta/g organic carbon to protect aquatic organisms 

from chronic effects. Total organic carbon in sediments 

under the Cougar Smith Bridge was calculated at 1.04%. At 

this level of TOC, the New York State benchmark would be 

420 μg penta/kg dry sediment weight. The maximum level 

observed was 4.5% of this benchmark, and adverse biological effects were not anticipated. 

Biological Response—Infauna 

The Satsop River supported a complex and abundant mix of 

megafauna that were not quantitatively sampled. These included numerous schools of juvenile salmon and hundreds of 

sculpins (Family Cottidae) observed under the bridge along 

with dozens of crayfish (Pacifastacus sp.). 

Figure 25—Cougar Smith Bridge (a) as viewed from the 

north showing the treated-wood arches and (b) underside of bridge showing the treated-wood deck and 

support beams. 

Figure 26—West Fork Satsop River flowing under the 

Cougar Smith Bridge in southwestern Washington. 
55 

A total of 12,787 invertebrates in 69 taxa were identified in 

18 samples at the Cougar Smith Bridge. Taxonomic codes 

used in the database are defined in Appendix C. Dominant 

taxa were defined as those representing 1% of the total infaunal abundance (count = 128). The community was fairly 

evenly distributed, with 18 taxa meeting the 1% criterion. 

Dominant taxa are identified in Table 25. These 18 taxa 

(26% of total taxa) represented 90.2% of the total abundance. 

The dominant taxa included many recognized as intolerant of 

metals (particularly copper) and other pollutants (Orders 

Ephemeroptera, Plecoptera, and Trichoptera) and sensitive 

chironomids (including Tanytarsus sp. and Cricotopus sp.). 

The number of taxa, total species abundance, and dominant 

species abundance are displayed in Figure 27 as a function of 

distance upstream and downstream from the Cougar Smith 

Bridge. Figure 28 summarizes Shannon’s and Pielou’s 

Indexes at these same stations. 

Figures 27 and 28 indicate that all measures of biological 

response at distances downstream from the bridge were either 

equal to or exceeded those found at the upstream control. 

There was a significant increase in species richness (number 

of taxa) and diversity (Shannon’s Index) observed at those 

stations, ranging from under the bridge (0 distance) to 3 m 

(10 ft) downstream. As previously noted, penta levels were 

4.5% of the New York State sediment quality criterion for 

this compound, and penta was not observed in water. Biological effects were not expected at these low preservative 

levels, and adverse biological effects were not observed in 

association with this bridge. 

Laboratory Bioassays 

A series of 10-day amphipod (Hyalella azteca) tests were 

completed on sediments collected upstream and at distances 

of +0.45, +1.8, and +9.9 m (+1.5, +6.0, and +33.0 ft) downstream from the perimeter of the Cougar Smith Bridge 

(Table 26). 

The sediment grain size distribution in West Fork Satsop 

River sediments contained predominantly gravel and cobble. 

This distribution contained more coarse material than desired 

for Hyalella azteca bioassays. However, survival was generally good in all tests. Survival at the upstream control was 

93.8%, and this result provided an appropriate basis for 

judging toxicity associated with downstream treatment stations. Survival data was log transformed, and t-tests were 

used to evaluate the significance of differences in amphipod 

survival between stations. 

Penta was detected in sediments at the +0.45 and +1.8 m 

(+1.5 and +6.0 ft) downstream stations. Ten-day amphipod 

survival was not significantly reduced at either of these stations when compared with survival at the upstream control. 

Survival at the 9.9 m (33.0 ft) downstream station was significantly less than at the control, but penta was not detected 

in these sediments and the results are likely associated with 

some other factor. 

Risk Assessment Summary 

Creosote-treated bridges discussed previously in this report 

were sited over a slow-flowing stream carrying a large 

 

Table 24—Physicochemical characteristics observed at the Cougar Smith Bridge on the West Fork Satsop 

River in southwest Washington State on August 4, 1998a 

Station 

Depth 

(cm) RPD (cm) 

Pentab

 

(μg/kg) TVS (%) 

Sand 

(%) 

Silt and Clay 

(%) 

−9.9 m (−33 ft) upstream control 16.0 >10.0 7.5 U 1.56 5.99 1.86 

0.0 under bridge 25.3 >10.0 7.4 U 2.00 8.52 1.90 

0.45 m (1.5 ft) downstream 30.0 >10.0 9.0 1.91 1.06 1.69 

0.9 m (3.0 ft) downstream 28.0 >10.0 1.44 8.27 3.61 

1.8 m (6.0 ft) downstream 22.0 >10.0 6.6 J 1.89 11.21 4.02 

3 m (10 ft) downstream 15.5 >10.0 1.70 14.35 4.03 

6 m (20 ft) downstream 15.5 >10.0 7.4 U 1.66 9.25 5.19 

9.9 m (33 ft) downstream 16.0 >10.0 7.2 U 1.73 9.58 3.83 

a

All values are the mean of two replicates except the −9.9 m (−33 ft) upstream control and samples taken under 

 the bridge, which are the mean of three replicates. 

b

J indicates an estimated value when that result is less than the calculated detection limit. U indicates 

 the compound was not detected at the given detection limit. 
56 

bedload of sand and silt and clay. The invertebrate community associated with Pipe Creek was dominated by annelids 

and robust species of chironomids—both tolerant of the 

naturally stressful conditions found in this environment. The 

levels of PAHs found in sediments at Pipe Creek exceeded 

the TEL and PEL values for several compounds. However, 

adverse effects on the invertebrate community or in laboratory bioassays were not observed. It was hypothesized that 

adverse effects would be associated with these PAH concentrations found in more sensitive lotic communities that generally are found in faster flowing water with stable substrates. 

It was also hypothesized that at speeds greater than 20 to 25 

cm/s, where these more sensitive communities were likely to 

be found, that preservatives lost from bridge structures would 

be greatly diluted. 

Figure 27—The number of taxa, the abundance of all 

taxa, and only dominant taxa observed in 0.0309-m2

samples collected in the vicinity of the Cougar Smith 

Bridge crossing the West Fork Satsop River, in 

Washington State, August 5, 1998. 

Figure 28—Shannon’s and Pielou’s Indexes plotted 

against sediment penta levels as a function of distance 

upstream and downstream from the Cougar Smith 

Bridge. 

The Cougar Smith Bridge presents an example of this second 

type of environment. Current speeds in the West Fork Satsop 

River were moderately fast at >27 cm/s, and the bottom was 

very stable, being composed of primarily sand, gravel, and 

cobble. A diverse community of pollution-sensitive species 

was observed under and downstream from this bridge. Penta 

was not detected in the water and was detected at very low 

levels in sediments in only two samples. In neither case did 

the sediment levels reach even 5% of an appropriate chronic 

benchmark. Adverse biological effects were not anticipated, 

and none was observed. In general, the invertebrate community under and immediately downstream from the bridge was 

more abundant and diverse than was found either upstream or 

downstream at a distance of 9.9 m (33 ft). 

Table 25—Dominant invertebrate taxa observed in sediment samples collected in the vicinity of Cougar Smith 

Bridge crossing the West Fork Satsop River in Mason 

County, Washingtona

 

Dominant taxa Code 

Found 

in number of 

samples 

Total 

abundance 

Phylum Annelida ANNE 18 757 

Order Acarina AACAR 18 801 

Order Ephemeroptera 

Cynigma sp. AICYNA 17 157 

Baetis sp. AIBAESS 15 354 

Order Plecoptera 

Family Nemouridae AINEM 14 295 

Sweltsa sp. AISWE 18 1,899 

Order Coleoptera 

Cleptelmis sp. AICLEP 18 531 

Order Trichoptera 

Glossosoma sp. AIGLOS 12 154 

Brachycentrus sp. AIBRAS 12 224 

Apatania sp. AIAPAS 16 212 

Order Diptera 

Pedicia sp. AIPEOS 16 252 

Pericoma sp. ACPERS 15 147 

Family Chironomidae AICHR 18 454 

Trissopelopia sp. AICTRIS 18 246 

Synendotendipes sp. AISYET 18 493 

Tanytarsus sp. AITYS 18 3,539 

Cricotopus sp. AICCRIC 17 701 

Synorthocladius sp. ASYNS 16 313 

Total abundance of dominants 11,529 

a

Eighteen samples were collected, each covering 0.0309 m2. 

 

0

200

400

600

800

1000

1200

1400

-9.9 0 0.45 0.9 1.8 3.0 6.0 9.9

Distance from Cougar Smith Bridge (m)

Invertebrates (number/0.0309 m2

)

0

5

10

15

20

25

30

35

40

45

50

Dominant taxa abundance

Number of taxa

Total abundance

Number of taxa

Sediment levels of penta (

μg/kg)

0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

-9.9

Distance from Cougar Smith Bridge (m)

Shannon’s and Pielou’s Indexes

0

5

10

15

20

25

30

35

40

45

Pielou’s Index

Shannon’s Index

Penta

Penta benchmark

0 0.45 0.9 1.8 3.0 6.0 9.9
57 

Results From Upper Dairy Creek Bridge 

The Upper Dairy Creek Bridge (Oregon) was chosen as the 

second penta-treated bridge in this risk assessment (Fig. 29). 

The bridge carries a live load of HS20, including a 76.2-cm 

Table 26—Summary results of Cougar Smith Bridge 

sediment bioassays using the test amphipod Hyalella 

aztecaa

 

 Surviving amphipods (%) 

Replicate 

Laboratory 

control 

Upstream 

control 

0.45 m 

(+1.5 ft) 

1.8 m 

(+6.0 ft) 

9.9 m 

(+33.0 ft) 

1 100 100 90 90 60 

2 100 100 100 90 100 

3 100 100 90 100 80 

4 100 100 100 100 90 

5 100 90 100 90 60 

6 100 80 90 80 90 

7 100 80 90 100 70 

8 100 100 50 90 60 

Mean 100 93.8 88.8 92.5 76.3 

a Initial count of 20 per replicate. 

(3-in.) asphalt surface. It has a clear span of 17.4 m 

(58 ft) and sits on concrete abutments. All wooden components were treated to a retention of 96 kg/m3

 (0.6 lb/ft3) 

Type A penta. Figure 30 shows the general configuration of 

the bridge, and Figure 31 shows the wooden decking and 

support timbers. 

As shown in Figure 30, the Upper Dairy Creek was in a low 

to moderate flow state during the November 6, 1997, survey. 

The sampling transect represented the area of the stream bed 

with finest sediments and slowest flows. Current speeds 

along this transect were less than 3 cm/s. However, much 

faster currents were present throughout much of the stream 

channel, particularly in the middle of the stream where currents were >100 cm/s. 

The weather was overcast, and a light rain was falling. The 

water temperature was 10.8o

C at noon on December 7, 1997, 

and the air temperature was 14.5o

C. Water pH was essentially neutral, averaging 6.89 + 0.05 (three measurements). 

Water column alkalinity was 37.5 mg/L, and hardness was 

32.8 mg/L. Conventional physicochemical characteristics of 

sampled sediments are described in Table 27. Sediments 

along the sampled transect were dominated by sand and 

gravel, with minor amounts of silt and clay. 

Figure 29—Location of the Upper Dairy Creek Bridge in Washington County, Oregon. 

Upper Dairy Creek Bridge
58 

Sediment and Water Column Concentrations 

Penta was not detected in the water column at the upstream 

control station or under the bridge at the detection limit of 

0.25 μg/L (three samples each station) (Table 27). Penta was 

detected in all three samples collected under the bridge at 

concentrations ranging from 15 to 28 μg penta/kg dry 

sediment. A concentration of 17.0 μg/kg was also detected in 

sediments collected 0.9 m (3.0 ft) downstream from the 

downstream perimeter of the bridge. Penta was not detected 

in other sediment samples at this bridge. 

Sediment Benchmark 

As previously noted, the New York State Department of 

Environmental Conservation (NYSDEC 1993) has established a freshwater sediment criterion for penta at 

40 μg penta/g organic carbon to protect aquatic organisms 

from chronic effects. The mean of TVS in sediments from 

the vicinity of Upper Dairy Creek Bridge was 5.88 + 0.59, 

indicating a mean TOC content of about 3.53%. At this level 

of TOC, the New York State benchmark would be 1,412 μg 

penta/kg dry sediment weight. The maximum observed level 

(28 μg/kg) was 1.98% of this benchmark, and adverse biological effects were not anticipated in association with preservative loss from the bridge. 

Biological Response—Infauna 

A total of 9,158 invertebrates in 74 taxa were identified in 

18 samples at the Upper Dairy Creek Bridge. Taxonomic 

codes used in the database are defined in Appendix C. 

Dominant taxa were defined as those representing 1% of the 

total infaunal abundance (count = 92). A total of 16 taxa met 

the 1% criterion. Dominant taxa are identified in Table 28. 

These 16 taxa (22% of total taxa) represented 90.2% of the 

total abundance. 

Figure 30—Upper Dairy Creek Bridge (Washington 

County, Oregon, Bridge Number 1366) as it appeared 

November 6, 1997, during this biological evaluation. 

Figure 31—Bridge deck and support structure under 

the Upper Dairy Creek Bridge. 

Table 27—Physicochemical characteristics observed at the Upper Dairy Creek Bridge, in Washington 

County, Oregon, November 7, 1997a

 

Station 

Depth 

(cm) 

RPD 

(cm) 

Pentab

(μg/kg) TVS (%) Sand (%) 

Silt and 

clay (%) 

−9.9 m (−33 ft) upstream control 21.0 >10.0 9.7 U 5.1 + 0.7 21.9 + 9.6 5.2 + 2.9 

0.0 under bridge 40.0 >10.0 20.0 + 7.9 7.9 + 1.3 62.2 + 3.5 12.7 + 3.5 

0.45 m (1.5 ft) downstream 29.0 >10.0 11.0 U 5.1 57.4 13.5 

0.9 m (3.0 ft) downstream 31.0 >10.0 17.0 4.6 30.9 5.9 

1.8 m (6.0 ft) downstream 34.0 >10.0 10.0 U 5.3 3.9 1.6 

3 m (10 ft) downstream 33.0 >10.0 8.6 U 3.7 20.4 5.7 

6 m (20 ft) downstream 29.0 >10.0 9.5 U 6.3 10.6 3.2 

9.9 m (33 ft) downstream 24.0 >10.0 11.0 U 6.6 26.0 4.4 

a

The upstream control and samples taken under the bridge are the mean of three replicates. 

 Where appropriate, values include 95% confidence limits on the mean.

b

U indicates that the compound was not detected at the given detection limit. 
59 

Dominant taxa represented 93.5% of the total invertebrate 

abundance in these samples. The invertebrate community 

was dominated by sensitive taxa in the orders Ephemeroptera 

and Trichoptera. Annelids and chironomids made up only 

8.5% of the number of dominant taxa. The remaining 91.5% 

were members of the orders Ephemeroptera or Trichoptera. 

The EPT are recognized as sensitive indicators of pollution. 

The number of taxa, total species abundance, and dominant 

species abundance are displayed in Figure 32 as a function of 

distance upstream and downstream from the Upper Dairy 

Creek Bridge. Figure 33 summarizes Shannon’s and Pielou’s 

Indexes at these same stations. 

The sensitive species dominating the invertebrate community 

in this section of Upper Dairy Creek generally prefer hard 

and stable substrates. Figures 32 and 33 suggest that TVS 

and proportion sand, silt, and clay were higher at stations 0 

and 0.45 m (1.5 ft) than at other stations examined in this 

Figure 32—The number of taxa, the abundance of all 

taxa, and only dominant taxa compared with the 

percentage of sand, silt, and clay in sediments and 

penta concentrations observed in 0.0309-m2

 samples 

collected in the vicinity of the Upper Dairy Creek Bridge 

in Washington State, November 7, 1997. 

Figure 33—Shannon’s Index plotted against sediment 

physicochemical parameters, total volatile solids, and 

percentage of sand, silt and clay as a function of 

distance upstream and downstream from the Upper 

Dairy Creek Bridge. 

assessment. Table 29 summarizes the results of a series of 

t-tests on log transformed data comparing the upstream control with downstream treatment stations. Few significant 

differences were observed. The proportion of sand, silt, and 

clay was only significantly greater (α = 0.05) in sediments 

from under the bridge compared with the upstream control. 

Table 29 indicates that Shannon’s Index is the only biological metric that was significantly different in downstream 

samples compared with the upstream control. 

The maximum observed sediment concentration of penta was 

low, being less than 1.98% of the New York benchmark at 

the calculated level of organic carbon. The variation in 

Table 28—Dominant and subdominant invertebrate taxa 

observed in sediment samples collected in the vicinity 

of Bridge 146 crossing Pipe Creek in Cass County, 

Indianaa

 

Dominant taxa Code 

Found 

in number of 

samples 

Total 

abundance 

Phylum Annelida ANNE 16 137 

Order Ephemeroptera 

Ephemerella sp. AIEPHS 18 702 

Paraleptophlebia sp. AIPAR 15 167 

Epeorus sp. AIEPES 18 580 

 cf. Heptagenia sp. AICHEPS 15 202 

Rhithrogenia sp. AIRHIS 14 112 

cf. Baetis sp. AIBAE 17 313 

Zapada sp. AIZAPS 17 318 

Sweltsa sp. AISWE 14 93 

Family Perlodidae AIPER 15 93 

Order Trichoptera 

Glossosoma sp. AIGLOS 18 4,771 

cf. Hydropsyche sp. AIHYHS 18 194 

Family Brachycentridae AIBRAD 17 165 

Family Lepidostomatidae 

Lepidostoma sp. AILPS 17 127 

Order Diptera 

Family Chironomidae 

Polypedilum sp. AIPPS 11 279 

Eukiefferiella sp. AEUKS 18 313 

Total abundance of dominant taxa 8,566 

a

Eighteen samples were collected, each covering 0.0309 m2. 

0

100

200

300

400

500

600

700

Distance from Upper Dairy Creek Bridge (m)

Dominant taxa and total abundance

0

10

20

30

40

50

60

70

80

90

Penta (

μg/kg); number taxa; and 

sand, silt, and clay (%)

Dominant taxa abundance

Total abundance

Sand, silt & clay

Penta

Number of taxa

-9.9 0 0.45 0.9 1.8 3.0 6.0 9.9

800 100

0

30

60

90

120

150

Distance from Upper Dairy Creek Bridge (m) 

Penta concentrations (

μg/kg) 

0

1

2

3

4

5

6

7

8

9

10

Total volatile solids (%) and 

Shannon's Index 

Penta

Penta benchmark

Sand, silt, & clay

TVS

Shannon’s Index

-9.9 0 0.45 0.9 1.8 3.0 6.0 9.9
60 

sediment concentrations of penta observed under the bridge 

was high enough to avoid significance at the 0.05 level. 

However, at α = 0.055, the differences would be significant. 

The relationship between sediment physicochemical characteristics and biological response was explored using varimax 

normalized principle components analysis (Fig. 34). The first 

two factors, displayed in Figure 34, explained 47.5% of the 

variation. The next two factors explained an additional 

31.9%, giving a total of 79.42% for the four factors considered in this analysis. 

Factor 1 was significantly positively correlated with the 

mayflies (LAIEPHS = Ephemerella sp., LAIPAR = Paraleptophlebia sp., LAICHEPS = Heptagenia sp., and LAIZAPS 

= Zapada sp.). It was significantly negatively correlated with 

the percentage sediment sand, silt, and clay (LFINES). 

Factor 2 was significantly correlated with the cadisflies 

Glossosoma sp. (LAIGLOS), Family Brachycentridae 

(LAIBRAD), total abundance (LABUND), and dominant 

species abundance (LDABUND). Glossoma sp. was the most 

abundant invertebrate in this study; therefore, its abundance 

is highly correlated with the metrics total abundance (LABUND) and dominant species abundance (LDABUND). 

There were no significant negative correlations associated 

with factor 2. 

Factor 3 was significantly correlated only with the chironomid, Polypedilum sp. (LAIPPS). 

Factor 4 was significantly positively correlated with the 

mayflies Epeorus sp. (LAIEPES) and Baetis sp. (LAIBAE) 

and the caddisfly Hydropsyche sp. (LAIHYHS). Factor 4 was 

significantly negatively loaded with the physicochemical 

variable total volatile solids (TVS). Sediment concentrations 

of penta were not significantly correlated with any of the four 

factors. However, it was moderately negatively correlated 

(−0.47) with factor 1 and fairly strongly negatively correlated 

(−0.61) with factor 4. This is likely because factor 1 was 

negatively correlated with fines (sand, silt and clay) and 

factor 4 was negatively correlated with TVS. These last two 

physicochemical parameters increased in depositional environments where penta could also be expected to accumulate. 

This analysis suggests that the sensitive community of mayflies and cadisflies resident in Upper Dairy Creek may have 

been reduced in areas containing more sand, silt, and clay 

(LFINES) and increased concentrations of organic matter 

(TVS). Sediment concentrations of penta were well below 

the New York State chronic sediment quality criterion, and 

the position of LSEDPENT was probably due to its accumulation in depositional areas with increased fines and TVS. 

This analysis is presented in a more intuitive manner in 

Figure 35, which summarizes the results of a principal factors 

(MINRES) extraction followed by Varimax normalization 

using the same data. 

Factor 1 is defined significantly by the abundance of the 

mayflies Ephemerella sp. and Zapada sp., the chironomid 

Eukiefferiella sp., and Shannon’s Index in the positive direction, and by the proportion of fines (LFINES) in the negative 

direction. Significant correlations with factor 2 involve the 

abundance of invertebrates (Glossosoma sp., Family Brachycentridae, total abundance, and dominant taxa abundance). 

Three variable clusters are evident in Figure 35. Cluster A 

consists of the physicochemical variables, TVS, sand, silt and 

clay (LFINES), and sedimented penta (LSEDPENT). Increased values of these variables are associated with depositional environments that are obviously somewhat unfavorable 

to the sensitive taxa resident in this part of Upper Dairy 

Creek. 

Table 29—Statistical significance of biological and sediment physicochemical endpoints between the 

upstream control and downstream stations at the Upper Dairy Creek penta-treated bridgea 

 Under 

bridge 

0.45 m 

 (1.5 ft) 

0.9 m 

(3.0 ft) 

1.8 m 

(6.0 ft) 

3 m 

(10 ft) 

6 m 

(20 ft) 

9.9 m 

(33 ft) 

Number taxa 0.58 0.24 0.25 0.22 0.42 0.13 0.22 

Total invertebrate abundance 0.97 0.28 0.10 0.06 0.10 0.055 0.44 

Abundance of dominant taxa 0.90 0.24 0.11 0.06 0.11 0.07 0.44 

Shannon’s diversity index 0.01 0.00 0.06 0.21 0.03 0.43 0.91 

Sediment penta 0.055 1.00 

TVS 0.09 0.95 

Sum percent sand, silt, and clay 0.01 0.21 

a

The probability that the mean value at the downstream station equals the mean value at the upstream control 

 is provided for each endpoint. Values that are significantly different (α = 0.05) are in bold type. 
61 

Figure 34—Varimax normalized principal components analysis displaying the 

relationship between factors 1 and 2 in a four-factor analysis. Variable codes 

are given in Appendix B. 

Figure 35—Principal factors (MINRES) extraction and varimax normalization of 

independent physicochemical variables and biological response variables 

observed in sediments at the Upper Dairy Creek Bridge in Washington County, 

Oregon, November 7, 1997.

 

Factor 1

Factor 2

LAIEPHS

 

LAIPAR

 

LAIEPES

 

LAICHEPS

 LAIBAE

 LAIZAPS

 

LAIGLOS

 

LAIHYHS

 

LAIBRAD

 

LAIPPS

 LAEUKS

 

LTAXA

 

LABUND

LSHANNON

 

LSEDPENT

 

LTVS

 

LFINES

 

LDABUND

 

-0.8

-0.4

0.0

0.4

0.8

1.2

-1.0 -0.6 -0.2 0.2 0.6 1.0

Factor 1

Factor 2

LAIEPHS

LAIPAR

LAIEPES

LAICHEPS

LAIBAE

LAIZAPS

LAIGLOS

LAIHYHS

LAIBRAD

LAIPPS

LAEUKS

LTAXA

LABUND

LSHANNON

LSEDPENT

LTVS

LFINES

LDABUND

-0.6

-0.4

-0.2

0.0

0.2

0.4

0.6

0.8

1.0

1.2

-1.0 -0.6 -0.2 0.2 0.6 1.0 1.4

Cluster A 

Cluster B 

Cluster C 
62 

Cluster B represents species and biological metrics contributing to measures of invertebrate abundance. These variables included dominant species abundance (LDABUND), 

total abundance (LABUND), and the most abundant taxon 

Glassosoma sp. (LAIGLOS). Note that values of Shannon’s 

Index are far from and negatively correlated with Clusters A 

and B, as expected. Cluster C contains a group of taxa that 

are important in determining the total number of taxa 

(LTAXA) observed in Upper Dairy Creek sediments. In 

addition, taxa in cluster C appear to be unadapted to the 

depositional environments characterized by cluster A. 

Laboratory Bioassays 

The preceding analysis hypothesized that the low levels of 

penta observed in Dairy Creek sediments are not associated 

with adverse biological effects but rather that they are correlated with sediment fines and TVSs characteristic of depositional areas. In addition, it has been hypothesized that it is 

the increased fines and TVS found in depositional samples 

that are responsible for the minor decreases observed in some 

biological endpoints. 

A series of 10-day amphipod (Hyalella azteca) tests were 

completed on sediments collected upstream and at distances 

of 0, +1.8, +9.9 m (0, +6.0, and +33.0 ft) downstream from 

the downstream perimeter of the Cougar Smith Bridge. The 

results are summarized in Table 30. 

The sediment grain size distribution in Upper Dairy Creek 

sediments contained significant quantities of gravel and 

cobble but held more sand than those at the Cougar Smith 

Bridge. This distribution is still coarser than desired for 

Hyalella azteca bioassays. However, survival was excellent 

in all tests. Survival at the upstream control was 96.0%, and 

this result provides an appropriate basis for judging toxicity 

associated with downstream treatment stations. Survival data 

were transformed using the arcsin (square root (proportion 

surviving)) transformation, and t-tests were used to evaluate 

the significance of differences in amphipod survival between 

the upstream control and downstream treatment stations. 

Penta was detected in sediments at the +0 and 0.9 m (+0.0 

and +3.0 ft) downstream stations. Ten-day amphipod survival 

was not significantly reduced at any downstream station 

compared with survival in either laboratory control sediments 

or in sediments from the more appropriate upstream control 

station. These results support the previous hypothesis that 

there was no adverse biological response to the use of penta 

in treating the wooden components of the Upper Dairy Creek 

Bridge. 

Risk Assessment Summary 

Creosote-treated bridges discussed previously in this report 

were sited over a slow-flowing stream that carried a large 

bedload of sand and silt and clay. The invertebrate community associated with Pipe Creek was dominated by annelids 

and robust species of chironomids—both tolerant of the 

naturally stressful conditions found in that stream. The level 

of PAHs found in sediments at Pipe Creek exceeded the TEL 

and PEL values for several compounds. However, no adverse 

effects on the invertebrate community or in laboratory bioassays were observed. It was hypothesized that adverse effects 

would be associated with these PAH concentrations in more 

sensitive, lotic communities that generally are found in faster 

flowing water with stable substrates. It was also hypothesized 

that at current speeds greater than 20 to 25 cm/s, where these 

more sensitive communities were likely to be found, 

preservatives lost from bridge structures would be greatly 

diluted. 

The Dairy Creek Bridge presents an intermediate example 

lying between the fast flows found at the West Fork Satsop 

River Bridge and the slow flowing waters of Pipe Creek. 

Current speeds in the West Fork Satsop River were moderately fast (>27 cm/s) and the bottom was very stable, being 

composed of primarily gravel and cobble. Flows along the 

sampling transect at Upper Dairy Creek were significantly 

less (about 3 cm/s), particularly under the bridge. A diverse 

community of pollution-sensitive mayflies and cadisflies was 

observed under and downstream from the Upper Dairy Creek 

Bridge. Penta was not detected in the water and was detected 

at levels representing less than 2% of the New York State 

sediment quality criteria at two stations. 

Table 30—Summary results of Upper Dairy Creek Bridge 

sediment bioassays using the test amphipod Hyalella 

aztecaa 

 Surviving amphipods (%) 

Replicate 

Laboratory 

control 

Upstream 

control +0.0 

+1.8 m 

(+6 ft) 

+9.9 m 

(+33 ft) 

1 100 100 100 95 100 

2 100 95 85 95 100 

3 95 100 100 95 90 

4 100 95 100 95 100 

5 100 90 95 100 95 

Mean 99.0 96.0 96.0 96.0 97.0 

95% CI 1.96 3.67 5.71 1.96 3.92 

p1 b 1.00 0.331 0.498 0.070 0.221 

p2 c 0.331 1.00 0.671 0.810 0.679 

a

Initial count of 20 per replicate.

b

p1 is the probability that amphipod survival at the Dairy 

 Creek station is equal to survival in the laboratory control 

 sediment. 

c

p2 is the probability that amphipod survival in downstream 

 sediments at the Dairy Creek Bridge is equal to survival in 

 sediments collected at the upstream control station. 
63 

A small decrease in several biological endpoints, particularly 

Shannon’s Index, was observed directly under the Dairy 

Creek Bridge. This decrease appears associated with the 

significantly greater proportion of sand, silt, and clay found 

in under-bridge sediment samples than was collected downstream. This part of the study could be criticized for not 

collecting samples further into the channel under the bridge 

where the substrate was more characteristic of the upstream 

control and downstream treatment stations. However, it was 

the author’s opinion that collecting samples further in the 

channel would not represent a worst case because it was very 

unlikely that penta would have been detected in sediments 

underlying the swift currents. All samples collected in support of this study were taken from areas representing worstcase (that is, slow flow) conditions. 

The weight of evidence presented in this analysis, including 

the low levels of penta observed in sediments, lack of penta 

in water, lack of toxicity in laboratory bioassays, and the 

relationship between the larval insect community and sediment grain size and TVS all support the conclusion that no 

adverse biological effects were associated with the use of 

penta in constructing the Upper Dairy Creek Bridge. Note 

that the biological effects discussed in this analysis were 

minor and generally not statistically significant between the 

upstream control and treatment stations. There were no adverse effects to document on November 7, 1997. 

CCA-Treated Bridges 

Chromated copper arsenate (CCA) is a formulation of copper, chromium, and arsenic dissolved in an aqueous solution. 

CCA was first developed in 1933 and has been widely used 

throughout the world as a wood preservative for 60 years. 

The metals in CCA-treated wood undergo chemical change 

after impregnation, resulting in their being bonded or fixed to 

wood fibers. These formulations combine the fungicidal 

properties of copper with the insecticidal properties of arsenic pentoxide. In CCA, the fixation of arsenic and copper is 

dependent on the presence of chromium. Several formulations, using different forms of the metals and marketed under 

a variety of trade names, are available. The American Wood 

Preservers’ Association (AWPA) classifies these formulations as CCA Types A, B, and C. Each type may exhibit 

slightly different behavior in aquatic environments. However, 

these differences are small and only the most commonly used 

formulation of CCA was considered in this review. 

Ammoniacal copper arsenate (ACA) has recently been improved for the replacement of some of the arsenic with zinc. 

The current formulation of ammoniacal copper zinc arsenate 

(ACZA) contains 25% zinc as zinc oxide, 50% copper as 

cupric oxide, and 25% arsenic as arsenic pentoxide. Ammonia in this formulation catalyzes the fixation of copper, arsenic, and zinc to the wood fibers (Lebow and Morrell 1993). 

Both arsenic and chromium are heavy metal poisons. Both 

have chronic and acute environmental health risks associated 

with them. Copper does not generally constitute a human 

health risk. However, low concentrations of copper, in 

certain ionic forms, are highly toxic to fauna and flora. 

The known toxicity to humans of arsenic and chromium 

has resulted in concern about introducing them into the 

environment. 

Several reviews assessing the environmental risks associated 

with treated wood have been compiled (Konasewich and 

Henning 1988, Ruddick and Ruddick 1992, Stranks 1976, 

and USDA 1980). The conclusion reached in these papers 

was that the use of treated wood causes no significant hazard 

to the environment. However, all these reviews suffer from 

lack of quantitative analysis, leaving some doubt about the 

risks associated with using treated wood in aquatic environments. Cooper (1990) discusses the considerable confusion 

created by contradictions in the technical literature regarding 

the leaching of metals from CCA-treated wood products. 

Brooks (2000) reviewed the toxicity of copper, chromium, 

and arsenic in freshwater ecosystems. The following is 

intended to expand that discussion to include marine 

environments. 

Background Levels and Sources 

The metals used in wood preservatives are naturally occurring elements. The purpose of this discussion is to gain an 

understanding of the background concentrations of these 

elements and the natural and anthropogenic sources that 

contribute to present environmental levels. 

Arsenic 

Arsenic is found in rocks and soil from concentrations of less 

than 1.0 to several hundred mg/kg. Andreae (1978) found 

background levels of arsenite (arsenic(III)) to be less than 

0.9 μg/L in seawater and total inorganic arsenic (primarily 

arsenate or arsenic V) to be less than 1.5 μg/L. Neff (1997) 

reported that the average concentration of total arsenic in the 

ocean is about 1.7 μg/L. He found that uncontaminated nearshore sediments contained between 5 and 15 μg As/g dry 

sediment weight. Arsenic concentrations in deep-sea sediments were greater, at about 40 μg/g1

. Contaminated sediments in urban and industrial areas may contain several 

thousand micrograms arsenic per gram of dry sediment. 

Finally, Neff (1997) reported that river sediments in England 

contained between 7 and 950 μg As/g dry sediment. Penrose 

and Woolson (1974) reported arsenic levels of 2.3 to 

8.3 μg/L in the Caribbean and 0.8 to 4.5 μg/Lin the Gulf of 

Mexico. The U.S. EPA (USEPA 1985) reported arsenic 

levels of 1.5 μg/L in Puget Sound water. Brooks (unpublished data) found arsenic levels of 19 μg/L in marine water 

samples in Little Skookum Inlet, South Puget Sound. There 

is a consensus among other authors (Waslenchuk 1977, 1978, 
64 

Sanders and Windom 1980) that total arsenic concentrations 

in marine waters generally range between 1 and 1.5 μg/L. 

Andreae (1978) found significantly more reduced arsenic 

(arsenite) than would be expected in highly oxygenated water 

where chemical equilibrium models suggest that most of the 

arsenic should be in the less toxic arsenate form. The ratio of 

arsenite to arsenate is correlated with chlorophyll α production, suggesting that the speciation of arsenic in natural waters is highly influenced by biological activity. Sanders and 

Windom (1980) estimated that as much as 15% to 20% of 

total arsenic is reduced to arsenite by phytoplankton during 

spring and fall blooms on the continental shelves. 

Woolson (1983) suggested that the natural arsenic cycle is 

not greatly disturbed, on a global basis, by human activities. 

He reported that volcanoes constitute the major natural 

source and estimated their contribution at 70,000 metric 

tons/year. All other natural sources contribute only 8,000 

metric tons/year. Anthropogenic sources add nearly 25% to 

the total global loading. Iron, steel, lead, zinc, and copper 

production contribute 82% of the anthropogenic emissions of 

23.6 × 109

 g/year. Pesticides account for 0.20 × 109 g/year, 

or less than 1% of total anthropogenic input. Neff (1997) 

reported total natural emissions of arsenic to the atmosphere 

at about 45,000 metric tons/year and anthropogenic emissions at 28,000 metric tons/year for a total atmospheric loading of 73 metric tons/year. 

Woolson discussed the arsenic cycle and suggested that the 

final environmental fate of all arsenic is incorporation into 

oceanic sediments. Characteristic levels of arsenic in Washington State marine sediments are presented in Table 31. 

Values range from less than 10 mg/kg in reference areas to 

greater than 70 mg/kg in highly impacted areas like Elliott 

Bay. Carpenter (Penrose and Woolson 1974) reported arsenic levels of 3 to 15 mg/kg in sediments of Puget Sound. He 

also reported levels of 290 to 980 mg/kg in sediments near a 

Puget Sound smelter. 

The USDA (1980) reported extreme variation in the amount 

of arsenic found in fresh surface and ground waters of the 

world (undetectable to 276,000 μg/L). Table 32 provides a 

cross section of these data. Whanger and others (1977) found 

similar levels in Oregon waters. Oregon spring water typically contained 133 to 900 μg/L arsenic, and lake water was 

significantly less (<1 to 9 μg/L). 

These data suggest that on a global scale, little arsenic is 

contributed to aquatic environments by pesticide use. However, local arsenic levels may be highly influenced by anthropogenic inputs. Typical arsenic levels in marine water are 

in the low microgram per liter range. The report herein assumes that marine levels of total arsenic are 1.7 μg/L and that 

the ratio of arsenite to arsenate is approximately 1:4. Total 

arsenic concentrations of 10 to 30 mg As/kg appear reasonable for sediments in reference areas. Additional analysis in 

this report assumes that sedimented levels in industrial areas 

are about 100 mg/kg. Arsenic levels in freshwater appear to 

be greater than in marine waters. For purposes of additional 

analysis, an average total arsenic loading of 50 μg/L is assumed for lotic and lentic fresh waters. 

This review suggests that the state of arsenic in aquatic environments is controlled by a variety of parameters that are 

independent of the presence of treated wood. Therefore, the 

form of arsenic leached from treated wood probably has little 

effect on species of arsenic found in the water. These processes are rapid with first-order rate constants for arsenite 

oxidation reported in the range of 0.5 to 2.2/day. This analysis suggests that regardless of the arsenic species released 

from treated wood (most is in the form of arsenate), between 

1% and 20% will be converted to arsenite and between 80% 

and 99% will be in the form of arsenate. 

Chromium 

Chromium is the 21st most abundant element in the earth’s 

crust, with a mean concentration in U.S. soils of about 

40 mg/kg (Barnhart 1997). Eisler (1986) reviewed chromium 

Table 31—Metal levels in sediments of Puget Sounda 

Area 

Arsenic 

(ppm) 

Copper 

(ppm) 

Straight of Georgia <10 <50 

San Juan Islands <10 <50 

Bellingham Bay <30 >350 

Strait of Juan De Fuca <10 <50 

Penn Cove <10 <50 

Everett <70 >350 

Dyes Inlet <70 <150 

Elliott Bay >70 >350 

a

Puget Sound Environmental Atlas (1992) 

Table 32—Typical arsenic levels observed in fresh water 

Location 

Arsenic level 

(μg total arsenic/L) 

Glacial ice in Sweden 2.0 to 3.8 

Thermal waters in western 

United States 

20.0 to 3,800 

Columbia River, Washington State 0.2 to 86.9 

Yellowstone River 4.5 

California well water 10.0 to 2,000 

Washington well water 5.0 to 6.0 

Oregon well water 0.0 to 1,700 

California lakes 0.0 to 100 

Wisconsin lakes 4.0 to 117 

English rivers 7.0 to 950 
65 

hazards to aquatic species. He reported that the earth’s crust 

contains a mean of 125 mg/kg chromium. Natural-weathering 

processes release an estimated 32,000 metric tons of chromium per year. The annual world production of chromium is 

estimated at 7 × 106

 metric tons, and industrial inputs can be 

significant. Storm water, sewage treatment plants, metal 

plating, leather tanning, and mining industries contribute 

significant amounts of chromium to aquatic environments. 

Untreated industrial effluents contain up to 5 × 106 μg/L, and 

electroplating waste streams contain up to 1.29 × 106 μg/L. 

Background levels of 0.0 to 5 μg/L chromium were reported 

by Eisler (1986) for uncontaminated seawater. Freshwater 

levels were somewhat higher at 1.0 to 10 μg/L. The U.S. 

EPA (USEPA 1983) summarized chromium levels in surface 

waters of the United States. Values ranged from undetectable 

in some California well water to 2,790 μg/L for water near a 

cooling tower. Average values were between 4 and 20 μg/L

with a range of <1 to 112 for U.S. surface waters. The EPA 

did not list chromium as a metal of concern in Puget Sound. 

Turekian and Scott (1967) reported suspended loads of 

chromium in North American rivers ranging from 37,000 to 

460,000 μg/L. They estimated that 870 tons of chromium are 

transported by the Susquehanna River each year. Rivers 

located east of the Mississippi have higher concentrations of 

many metals than do western rivers. However, all these lotic 

systems transport significant amounts of chromium to the 

estuaries at their mouths. Eisler (1986) reported that marsh 

sediments receiving fertilizers containing sewage sludge for 

7 years contained between 2,150 and 4,750 mg/kg of chromium. Sediments associated with industrial outfalls had as 

much as 54,300 mg/kg of chromium. Sediments from reference areas contained 50 to 54 mg/kg of chromium. 

In summary, uncontaminated seawater contains very low 

levels of chromium. This review finds that anthropogenic 

inputs of chromium are significant. This is particularly true in 

localized areas near urban or industrial centers because the 

metal is readily incorporated into sediments. Additional 

analysis in this report assumes a background level of 3 μg/L 

in marine water and 12 μg/L in freshwater. Based on values 

in the literature, sediments in unpolluted areas are assumed to 

have 52 mg/kg chromium and industrialized areas are assumed to have a level of 1,000 mg/kg. 

Copper 

Copper levels of 1 to 10 μg/L were reported by Boyle (1979) 

from unpolluted waters of the United States. However, concentrations downstream of municipal and industrial outfalls 

may be much higher (Hutchinson 1979). The U.S. EPA 

(USEPA 1985) reported low levels (0.25 μg/L) of copper in 

the waters of Puget Sound. 

Copper levels in Washington sediments (Puget Sound Environmental Atlas 1992) are summarized in Table 32. Lu and 

Chen (1976) reported reference area sediment copper levels 

in San Pedro Bay at 5 to 10 mg/kg. 

Based on available information, this report assumes that 

background water column levels of copper are <2.0 μg/L in 

marine environments and 2 μg/L in lotic and lentic freshwaters. Values of 10 to 30 mg/kg are assumed for unpolluted 

sediments and 350 mg/kg in sediments near industrialized 

areas. 

Cycling and Fate 

Arsenic 

The chemistry of arsenic in water is complex, and the form 

present in solution is dependent on such environmental parameters as pH, organic content, suspended solids, and sediment characteristics. Neff (1997) provides the most recent 

review of the chemistry of arsenic in aquatic environments. 

Thermodynamic considerations predict that at neutral pH and 

relatively high levels of dissolved oxygen, most arsenic 

should be oxidized to arsenate. Johnson (1972) found that 

marine bacteria can reduce arsenate to arsenite. Neff (1997) 

noted this same pathway but found that in aerobic water, 

arsenite is oxidized rapidly to the less toxic arsenate form, 

both abiotically and by bacteria. When oxygen is deficient, as 

may occur in association with the upwelling of deep ocean 

water, arsenite concentrations can significantly increase. 

However, arsenate is always the most abundant form of 

arsenic in oxygenated, biologically productive waters of 

the ocean. 

Onishi (Andreae 1978) reports that arsenic(III) (arsenite) 

represents only about 20% of the total arsenic found in 

seawater. Neff (1997) concluded that arsenite (the toxic 

form of arsenic) usually represents less than 1% to no 

more than about 10% to 20% of the total arsenic in marine 

environments. 

In addition to inorganic arsenic, several authors (Andreae 

1978, Johnson 1972, Lunde 1977, Penrose and others 1977) 

have demonstrated that bacteria, phytoplankton, marine 

invertebrates, and vertebrates can biotransform arsenic into 

relatively less toxic organic compounds. These reactions 

involve methylation and reduction to produce methylarsinic 

acid and dimethylarsinic acid. The low toxicity of these 

organic compounds allows high body burdens of arsenic in 

healthy organisms, which are eventually incorporated into the 

sediments. However, significant amounts of arsenate may be 

regenerated in the water column from phytoplankton that 

sink below the photic zone and perish. Thus, there is an 

arsenic cycle, which involves a cycling of arsenic through its 

various inorganic and organic forms. The relatively high 

levels of arsenic found in sediments, compared with levels in 

the water column, suggest that the ultimate fate of arsenic in 

aquatic environments is incorporation into sediments. 
66 

Chromium 

Two species of chromium are prevalent in aquatic environments: chromium(III) is less toxic than chromium(VI). 

Most chromium(VI) found in nature is a result of domestic 

and industrial emissions (Steven and others 1976). Interaction 

of chromium(VI) molecules with organic compounds can 

result in reduction to a comparatively less toxic trivalent 

form. However, in aerobic marine environments, chromium(VI) is the more abundant species. Chromate, hydrochromate, and dichromate are soluble in water and are therefore mobile in aquatic environments. 

The ultimate fate of chromium(VI) appears to be incorporation into fine-grained sediments with high organic and iron 

content. Adsorption of chromium(VI) onto sediments is 

dependent on salinity and is greatest at salinity of 0.1 to 

1.0 parts per thousand (Mayer and Schick 1981). However, 

its fairly high solubility allows easy migration in and out of 

the water column over aerobic sediments. Observed concentrations in European estuaries ranged from 3.9 mg/kg in 

intertidal sands to 162 mg/kg in anaerobic mud (Rehm and 

others 1984). 

Chromium(III) forms stable complexes with negatively 

charged inorganic and organic compounds. It is rarely found 

uncomplexed in waters, with decaying plant or animal tissues, or with silt and clay particles. Precipitated chromium 

hydroxides remain in the sediments under aerobic conditions. 

With low pH and anoxic conditions, chromium(III) hydroxides may solubilize as ionic chromium(III). However, Lu and 

Chen (1976) found that chromium was not significantly 

released from sediments into seawater under either oxidizing 

or reducing conditions. 

Copper 

Copper is found in natural water as a free ion or complexed 

with humic acids, carbonate, and other inorganic and organic 

molecules. Copper is an essential element in the normal 

metabolism of both plants and animals. Therefore, a significant portion of the copper found in both fresh and marine 

systems may be taken up by the biota. The ultimate fate of 

much of this copper is sedimentation. 

Harrison and others (1987) found very low copper levels 

(<12 mg/kg) in sandy substrates associated with power plant 

effluents. He suggested that the lack of organic matter in 

these sediments was responsible for the low copper content. 

Clarke (1974) noted that iron sulfide renders copper insoluble in anaerobic sediments. These reports suggest that copper 

accumulation in sediments is highly influenced by sediment 

chemistry and physical characteristics. Fine sediments, coupled with poor water circulation, could be expected to accumulate more copper than coarse sediments in highly oxygenated areas. Copper accumulations in fine-grained, anaerobic 

sediments are probably not biologically available. Therefore, 

these environments may serve as an important mechanism for 

the removal of excess copper from aquatic environments. 

The cycling of copper from sediments to the water column is 

a function of the sediment reduction–oxidation potential. Lu 

and Chen (1976) examined the release of copper from sediments as a function of sediment grain size and oxygen availability. Three oxidizing conditions were examined: oxidizing, 5 to 8 ppm dissolved oxygen; slightly oxidizing, <1 ppm 

dissolved oxygen; and reducing, S(-II)T = 15 to 30 mg/kg. 

Small amounts of bound copper were released to the overlying water in the reducing and slightly oxidizing environments, resulting in concentrations of 0.2 to 0.5 μg/L. Copper 

releases in the oxidizing environment resulted in significantly 

higher interfacial seawater concentrations (3.2 μg/L). This 

effect was slightly more pronounced in the coarsest sediment 

tested (silty-sand sediment). These data imply higher copper 

releases from sediments in aerobic (healthy) environments. 

There are two ways to look at this phenomenon. 

In more coarse-grained, highly oxygenated sediments, bound 

copper is more easily lost to the water column and dispersed 

over greater distances, until the copper finds anaerobic sediments, where it will likely be buried and eventually incorporated into the lithosphere. These anaerobic sediments support 

reduced infaunal and epifaunal communities of organisms. 

Therefore, we would expect reduced environmental impacts 

from copper incorporated into anaerobic sediments. 

In enclosed bodies of water with coarse grained, aerobic 

sediments, this study suggests that copper will not be as 

tightly bound to the sediments and will migrate into the 

interstitial and surficial water where it is bioavailable. Data 

were not provided on the copper species released from the 

sediments; therefore, it was difficult to assess the toxicity of 

the released copper. 

The work of Lu and Chen (1976) suggests that caution is 

appropriate when dealing with copper materials in poorly 

flushed embayments with aerobic (>2 to 3 mg/L dissolved 

oxygen) sediments. These arguments suggest that anaerobic 

sediments are a more efficient trap for released copper and 

that reduced environmental risks should be anticipated from 

copper releases associated with anaerobic sediments compared with those associated with aerobic sediments. 

Data presented in Lu and Chen (1976) are not sufficient to 

develop an expression describing copper releases from sediments at a variety of sediment concentrations. These effects 

appear to be subtle, and their exclusion should not significantly flaw the risk assessment. This discussion is provided 

as background for project proponents and permit writers. 

Consideration of these factors may be important when considering the relative risks associated with different wood 

treatments. 
67 

Bioconcentration, Bioaccumulation, 

and Biomagnification 

Chromium and copper are essential micronutrients for plants 

and animals. Their uptake and metabolism is a normal biological process. The pentavalent form of arsenic (arsenate) is 

chemically similar to phosphate, and arsenate may be readily 

taken up by plants and animals in their efforts to sequester 

phosphate for normal cellular metabolism. This section 

discusses the potential for the bioaccumulation of each of 

these metals by aquatic plants and animals. 

Arsenic 

Arsenic(III) is a potent toxicant in mammals (including 

humans), and there are considerable data describing its 

bioaccumulation. 

Penrose and others (1977) examined the arsenic budget in a 

sea urchin–alga system and concluded that organic arsenic is 

rapidly excreted by most organisms. Therefore, while there 

may be significant bioconcentration of arsenic from surrounding waters, there is no apparent biomagnification in 

food chains. Organisms containing high levels of arsenic in 

their tissues tend to be those that are prone to incidental 

ingestion of sediment particles while feeding. 

Arsenic concentration from ambient water was also reported 

by Schroeder and Balassa (1966), Lunde (1970, 1972), and 

Fowler and others (1975). High levels of arsenic in marine 

animals were reported by USDA (1980) from around the 

world. Reported levels of arsenic, expressed as a proportion 

of wet tissue weight, for some typical marine species are 

provided in Table 33, excerpted from USDA (1980). Woolson (1975) reported that arsenic concentrations are 10 to 100 

times greater in marine fish and shellfish than in freshwater 

species. However, as seen in previous sections, reported 

arsenic concentrations in marine waters are typically less 

(1.5 μg/L) than in freshwater (see Table 32). No plausible 

explanation for this apparent contradiction was offered in the 

literature. 

Penrose and Woolson (1974) reviewed studies by Fernandez 

del Riego, Seydel, and Lunde, who suggest that arsenic is not 

biomagnified in food chains. Boothe, Knauer, Black, and 

Penrose (Penrose and Woolson 1974) suggested that marine 

organisms rapidly excrete arsenic ingested in food. Woolson 

(1975) summarizes his review of arsenical bioaccumulation 

by noting that 

Arsenic is bioconcentrated by aquatic organisms but not 

biomagnified. Plants usually accumulate more arsenic 

than fish, and Crustacea accumulate intermediate 

amounts. Marine organisms normally contain more arsenic than their freshwater counterparts. However, the 

arsenic contained in the organisms is apparently not 

toxic to animals or humans, and is readily excreted. 

(Woolson 1975) 

Neff (1997) reviewed the literature describing the bioaccumulation of arsenic in marine environments and humans 

and noted that inorganic arsenic is poorly bioconcentrated 

from water but that organic arsenic compounds have much 

higher BCFs. The inorganic arsenic that is bioconcentrated is 

rapidly converted to organoarsenic compounds, which are 

not toxic. Arsenobetaine is the most abundant form of arsenic 

in tissues of most marine animals examined to date. It usually 

represents 50% to more than 95% of the total arsenic in 

crustaceans and fish. Neff (1997) concluded that inorganic 

arsenic, at concentrations typically found in productive, 

oxygenated seawater and brackish waters, does not represent 

a significant hazard to marine organisms and ecosystems. 

The available evidence indicates that arsenic does not biomagnify in food chains. It appears that arsenic ingested at 

lower levels of the food web is converted to organic molecules, which are rapidly excreted at the next trophic level. 

For purposes of the analysis in this report, it will be assumed 

that levels of arsenic are dependent on ambient water levels 

and they are not biomagnified in higher trophic levels. 

Chromium 

Eisler (1986) reported that algae and higher plants accumulate chromium from seawater by factors of up to 8,600 and 

from solutions containing 50 mg/L chromium by a factor of 

18 in 48 h. Although chromium is abundant in primary producers, there is little evidence of biomagnification through 

marine food chains. Baptist and Lewis (Eisler 1986) followed the transfer of chromium through an experimental 

food chain and observed a decline in the concentration of 

chromium through each of four trophic levels. A comparison 

of the results of this food chain study with measurements of 

direct chromium uptake from seawater suggests that direct 

uptake is a far more important pathway than assimilation 

Table 33—Arsenic content of aquatic animal life 

Arsenic (μg arsenic/g 

dry tissue weight) 

Marine 

Crab 27.0 to 52.5 

Clams (all species) 90.0 to 127.2 

Oysters (Crassostrea virginica) 0.6 to 42.8 

Lobster (Panulirus borealis) 0.003 to 9.6 

Tuna 0.7 to 4.6 

Fresh water 

Trout 0.069 to 0.149 

Perch (Perca fluviatilis) 0.6 

Bass (Micropterus salmoides) 0.070 to 0.930 

Channel catfish (Ictalurus 

punctatus) 

0.0 to 3.100 
68 

through the food chain. Bioconcentration factors (BCFs) for 

numerous aquatic species are given by U.S. EPA (USEPA 

1983). The reported BCF for chromium(VI) in fish muscle is 

less than 1.0. EPA obtained BCF values of 125 and 192 for 

chromium(VI) in oysters and blue mussels. The EPA document also gives values for chromium(III) and concludes that 

they are similar to those given for chromium(VI). The EPA 

conclusion was that mean BCF values of 0.5 and 130 are 

appropriate for fish muscle and bivalve mollusks, respectively. These are both relatively low BCFs. It appears that 

chromium is not biomagnified in food chains and that chromium concentration levels at all trophic levels are primarily a 

function of background levels in the water. 

Copper 

The National Academy of Sciences (NAS 1971) provides 

copper BCFs for numerous taxa. These values range from 

100× for benthic algae to 30,000× for phytoplankton. Mollusks concentrate copper by a factor of 5,000 in muscle and 

soft parts. No information was reviewed on the bioaccumulation of either copper or zinc by aquatic organisms. For the 

purposes of this paper, it is assumed that copper and zinc 

accumulation in aquatic organisms is primarily a function of 

metal concentration in the ambient water, and although many 

organisms may bioconcentrate copper, there is inadequate 

information describing the biomagnification of copper 

through food webs. 

Toxicity to Aquatic Fauna and Flora 

To assess the potential impact of CCA-treated wood used in 

aquatic environments, it is necessary to determine the minimum levels of these metals causing acute or chronic stress in 

populations of marine organisms. 

Arsenic 

Arsenic is a common environmental metal whose toxic properties have been known for centuries. The toxicology of 

arsenic may be divided into three general areas: direct inhibition of cellular respiration, mutagenic effects, and hemolysis. 

Baroni and others (1963) and Penrose and Woolson (1974) 

noted that controlled attempts to attribute carcinogenic properties to the arsenicals failed. Ferm (1977) demonstrated the 

teratogenic nature of sodium arsenate injected into a variety 

of experimental animals. 

Arsenical toxicity is dependent on the oxidation state, chemical form, and route of exposure. In general, arsenic acids are 

least toxic, followed by inorganic arsenate, arsenoxides, 

inorganic arsenite, and the trivalent organic and inorganic 

arsines are the most toxic. 

Eisler (1988) reported acute toxicity levels for a variety of 

freshwater and marine plants and animals associated with 

several species of arsenic. Lethal concentrations, which 

killed 50% of the invertebrate test organisms (LC50), are 

provided in Table 34. In marine water, it appears that arsenic 

levels in excess of 200 μg/L may result in the mortality of 

juvenile Dungeness crab and an unspecified species of red 

algae. NTIS (1986) reported acute values of arsenic(III) for 

12 saltwater animals. The range of sensitivity was from 232 

to 16,030 μg/L. Chronic stress was observed at about half 

these values, or 116 μg/L. Arsenic(V) was less toxic for the 

two invertebrates examined, with acute values of 2,000 and 

3,000 μg/L. None of these animals was as sensitive to arsenic 

as were some algae, which showed toxic responses to either 

arsenic(III) or (V) at values as low as 19 μg/L. 

Eisler’s (1986) data suggest that in freshwater, arsenic levels 

associated with acute toxicity appear to be somewhat higher, 

in the neighborhood of 900 μg/L. NTIS (1986) reported 

acute toxicity associated with arsenic(III) in 16 species of 

freshwater animals. An acute value of 812 μg As/L was 

found for cladocerans. At the other end of the range, the 

acute level for a midge was 97,000 μg/L. From these papers, 

it appears reasonable to assume an LC50 of 800 μg/L for the 

more sensitive freshwater species. NTIS (1986) suggested 

that chronic stress is encountered by all freshwater species at 

about 21% of their acute values. For the most sensitive 

species, this value would be 168 μg/L total arsenic. 

This review indicates that arsenite can cause chronic stress in 

several marine animals at levels as low as 168 μg/L in freshwater systems and 230 μg/L in marine systems. For most 

aquatic organisms, arsenate is far less toxic. However, for the 

most sensitive marine algae, this review indicates no difference in toxicity between the two primary valence states of 

arsenic (+3 and +5) and toxicity thresholds, which may be as 

low as 19 μg/L. 

Table 34—Lethal levels of arsenite in fresh-water 

plants and animals 

Taxa Arsenite (As+3) 

Fresh water 

Algae 1,700 to 4,000 (LC100) 

Cladocerans 1,300 

Amphipod 960 (28-day LC100) 

Goldfish 490 (7-day LC50) 

Fish 15,000-35,000 

Marine water 

Red algae 300 (LC100) 

Copepods 510 

Dungeness crab 230 

Oyster (eggs) 7,500 (48 h LC50) 

Blue mussels 16,000 (LC100) 

Pink salmon 3,800 (10-day LC50) 

a

Unless specified, values are for the LC50 expressed 

 in μg As/L.
69 

Chromium 

The toxicity of chromium to aquatic species can vary by an 

order of magnitude, or more, depending on a variety of biological and physical factors. These include differences associated with species, age, developmental state, temperature, 

pH, salinity, length of exposure, and interaction with other 

contaminants. Chromium(VI) is most toxic to the developmental stages of aquatic species in soft, freshwater, with low 

pH. Eisler (1986) reported a 96-h LC50 of 200 μg/L for 

salmon fingerlings and 495 μg/L for rainbow trout (Oncorhynchus mykiss) eggs. Most species of fish tolerate 

>10,000 μg/L, and Bluegills (Lepomis macrochirus) demonstrated a high toxic threshold at 213,000 μg/L in water with 

hardness of 120 mg/kg calcium carbonate. 

In marine water, chromium(VI) toxicity also varies by orders 

of magnitude, depending on the taxa. The range is about the 

same as for freshwater species. Polychaetes and larval crabs 

(Callinectes sapidus) are the most susceptible organisms at 

200 and 320 μg/L, respectively. 

Copper 

Copper is an essential element for most living organisms. It 

is added at a concentration of 2.5 μg/L in Guillard’s Medium 

F/2 to seawater for the optimum culture of marine algae 

(Strathman 1987). At concentration levels slightly above 

those required as a micronutrient, copper can be highly toxic, 

especially to the larval stages of marine invertebrates. A 

single copper fitting in a seawater system may destroy most 

invertebrate embryos being cultured in the laboratory. 

U.S. EPA’s (USEPA 1984) Ambient Water Quality Criteria 

reports that copper toxicity in aquatic environments is related 

to the concentration of cupric (Cu2+) ions. The cupric ion is 

highly reactive and forms various copper complexes and 

precipitates, which are significantly less toxic than the cupric 

ion (Knezovich and others 1981). Sunda (1987) has proposed 

a basic mechanism to explain the observed relationship 

between free ion activities and the bioavailability of metals 

such as copper. He observed that the complexed species of 

copper are charged or polar and cannot pass directly across 

the lipid bilayer of the cell membrane. Thus, transport of 

these metals across the membrane would require that they 

interact with specific metal transport proteins. Because the 

free ion activity is a measure of the potential reactivity of a 

metal, it reflects the ability of that metal to interact with these 

transport proteins. The many chemical forms of copper in 

aquatic environments are maintained in a dynamic state of 

equilibrium that depends on salinity, temperature, pH, alkalinity, dissolved oxygen, sediment physicochemical characteristics, and the presence of other inorganic and organic 

molecules. 

The dual nature of copper as an essential trace element and a 

potential toxin at low concentrations demands that organisms 

strictly regulate copper at internal levels suitable for 

metabolic requirements. Roesijadi (1980) reported that copper is normally present at relatively high levels in the tissues 

of marine animals (>1.0 μg/g). Roesijadi (1980), Harrison 

and others (1987), and Harrison and Lam (1985) reviewed 

both the environmental detoxification of copper and the 

physiological detoxification of copper by Mytilus edulis, 

Protothaca staminea, Patella vulgata, Ostrea edulis, and 

Littorina littorea. Copper detoxification and metabolic regulation were associated with copper binding by low- and highmolecular-weight metallothionein-like proteins in the digestive gland and sequestering of copper in lysosomes. 

Costlow and Sanders (1987) used a metal-chelate buffer 

system to regulate the free ion concentration of copper in 

seawater. They exposed crab larvae to a range of free cupricion concentrations and monitored survival, duration of normal development, and growth. The authors reported significant reductions in growth correlated with copper accumulation and concluded that when exposed to cupric ion 

concentrations in seawater that are below ambient concentrations, crab larvae are able to regulate their bioconcentration 

of copper. At higher concentrations of the cupric ion, copper 

bioconcentration increases and larval growth was inhibited. 

Harrison and others (1987) reported that copper discharged 

from the San Onofre power plant cooling system was mostly 

in bound forms under normal operating conditions. Their 

study found sufficient organic ligands available in ambient 

seawater to complex most of the copper and expected little or 

no impact from the discharges. Harrison and others (1987) 

conducted copper bioassays on a number of aquatic invertebrate and vertebrate species. They found that Crassostrea 

gigas embryos were most sensitive (48-h LC50 =10 μg/L) and 

larval herring the least sensitive. The range of 48-h LC50

values for copper was 10 to 2,000 μg/L. Dinnel and others 

(1983) published the results of copper toxicity bioassays on 

various life stages of several marine organisms. They reported a very low LC50 (1.9 μg/L) for the sperm of the red 

sea urchin (Strongylocentrotus franciscana). This value 

seems suspect because it falls within the range normally 

expected in unpolluted seawater. Other values from the 

Dinnel and others (1983) study are presented in Table 35. 

Gametes and the embryos of marine organisms are most 

sensitive to copper. Based on the previous discussion regarding the metabolic regulation of copper, it seems reasonable to 

suggest that the susceptibility of embryos to even low copper 

concentrations is associated with their inability to regulate 

cellular exposure to the cupric ion. Copper concentrations 

maintained at levels low enough to protect embryos are 

sufficient to ensure that toxic effects are not imposed on 

larvae and adult organisms. With the exception of the sperm 

of the red sea urchin, environmental levels less than 6 μg/L 

appear reasonable for the protection of aquatic life in marine 

waters. In areas where red urchins spawn, additional restrictions should be considered. 
70 

Because of the variety of molecular structures containing 

copper in aquatic environments, and a lack of definitive 

information about their relative toxicity, no single analytical 

measurement is ideal for expressing copper concentrations 

with respect to their potential toxicity to aquatic life. Baldwin 

(1989) advises that active copper (operationally defined by 

acidifying the aqueous sample to pH 4 with nitric acid and 

measuring the concentration of copper that passes through a 

0.45-μm membrane filter) is probably the best available 

measurement. 

This review revealed few copper toxicity data that included 

an analysis of the form of copper used in the bioassay. Most 

toxicity data are reported on the basis of total or dissolved 

copper. If bioassays are conducted in distilled water with low 

complexing capacity, there is significant potential to overestimate the toxicity of copper in the natural environment. 

If 2-mg copper sulfate is added to a liter of distilled water, 

much of this may become available in its toxic cupric ion 

form. However, the same amount of copper added to organically rich estuarine water could result in only a small fraction 

being present in the toxic form, the majority of the copper 

being detoxified by adsorption to sediments and precipitation 

or complexation with organic molecules. These comments 

indicate the difficulty in accurately assessing the impact of 

copper in natural environments. However, because of the 

potential for detoxification, water quality criteria based on 

total copper will result in conservative criteria. 

Summary 

This review of metal toxicity to aquatic organisms clearly 

demonstrates that copper is the metal of most concern in 

association with CCA-treated wood used in aquatic environments. From a purely biological point of view, the cupric ion 

should be maintained at less than 6 μg/L in marine environments. No observed effect levels (NOEL) for copper in 

embryos were not found in this review. However, it seems 

reasonable that these levels would be at least half the 48 hLC50 level, or 3 μg/L. Chromium is less toxic, and susceptible species have LC50 > 200 μg/L. Arsenic, which is notoriously toxic to humans and other mammals in some valence 

states, is highly tolerated by marine animals at levels up to 

230 μg/L. However, the susceptibility of some species of red 

algae impose an upper limit of 19 μg/L in marine environments where they are present. The most susceptible freshwater species, examined to date, tolerate up to 168 μg/L arsenic. 

Most potentially toxic substances are regulated at the Federal 

and State Levels. Therefore, it would be presumptuous to 

suggest that these values should be adopted as appropriate 

benchmarks when considering the use of treated wood in 

aquatic environments. This review is intended to provide 

insight to the regulatory standards, which are discussed in the 

following section. 

Regulatory Benchmarks 

Numerous states have adopted U.S. EPA water quality criteria for metals. Washington State Administrative Code (1992) 

defines water quality standards for surface waters. The code 

states that toxic substances shall not be introduced above 

natural background levels in waters of the state, which have 

the potential either singularly or cumulatively to adversely 

affect characteristic water uses, cause acute or chronic toxicity to the most sensitive biota dependent upon those waters, 

or adversely affect public health, as determined by the Department of Ecology. Table 36 lists criteria established for 

the protection of aquatic life in Washington State. 

The U.S. EPA marine water quality criteria are presented in 

Table 37. Direct comparison of the U.S. EPA criteria for 

copper and zinc with the Washington State Standards is not 

appropriate because the latter are variable and depend on 

water hardness. Table 37 also compares U.S. EPA criteria 

with existing average metal concentrations in unpolluted 

areas of Puget Sound. 

With the exception of the reported copper toxicity to red sea 

urchin sperm, both the State and Federal criteria are consistent with biological considerations and reflect significant 

safety factors for the protection of aquatic life. 

Marine Sediment Standard 

Washington is the only jurisdiction that has developed 

Marine Sediment Quality Standards for metals (Washington 

State Administrative Code 1991). These standards are based 

on apparent effects thresholds (AETs). These sediment standards are summarized in Table 38 together with the TEL and 

Table 35—Total copper toxicity measured in 

controlled bioassays 

Taxa EC50 or LC50 (μg Cu/L)

Sperm 

Purple sea urchins 34.0 

Oysters 12.1 

Salmon 44.2

Embyro 

Purple sea urchins 6.3 

Oysters 6.1 

Mussels 21.0 to 35.0 

Larvae 

Crab zoea 95.7 

Squid 309.0 

Cabezon 95.3 

Adults 

Sand shrimp 898.5 

Shiner perch 417.7 

Coho salmon smolt 601.0 
71 

PEL developed by the Florida Department of Environmental 

Protection. These are marine sediment quality benchmarks 

reported in Jones and others (1997). They are not sediment 

quality criteria. British Columbia has recommended sediment 

criteria based on the mean of the TEL and PEL. That value is 

included in Table 38. The last value ((TEL + PEL)/2) was 

used in this assessment to predict adverse biological effects 

in marine environments. 

Freshwater Sediment Benchmarks 

Regulatory criteria have not been developed for metals in 

freshwater sediments. However, Ingersol and others (1996) 

published benchmarks for evaluating freshwater sediments 

based on amphipod and midge bioassays. Benchmarks provided by these authors are summarized in Table 39. 

Summary of Sources and Toxicity 

in Aquatic Environments 

The metals of concern when considering the use of CCAtreated wood in aquatic environments are arsenic, chromium, 

and copper. These metals are natural components of the 

earth’s crust and are found at various concentrations in both 

freshwater and marine environments. All three are essential 

micronutrients (Neff 1997). However, high concentrations of 

arsenic, chromium, and copper are known to be toxic. Arsenic and chromium are tolerated at moderately high levels by 

aquatic species. Buchanan and Solomon (1990) concluded 

that chromium and arsenic concentrations in CCA leachate 

were deemed sufficiently low as to have little or no effect on 

LC50 determinations. 

With respect to CCA-treated wood, copper is the most toxic 

metal to aquatic organisms, particularly in marine environments. Copper toxicity, associated with the uncomplexed 

cupric ion, is most detrimental to the early life stages of 

marine invertebrates. These stages are not readily visible and 

therefore this toxicity is not necessarily apparent to casual 

observation. 

This risk assessment used regulatory criteria for dissolved 

metals in freshwater and marine environments. The recommendations of Brooks (2000) and the mean of TEL and PEL 

was used in lieu of freshwater sediment quality criteria, 

which have not yet been adopted. The Washington State 

Sediment Quality Criteria and the mean of the TEL and PEL 

were used to evaluate marine sediment metal concentrations 

associated with CCA-treated timber bridges. 

Results From Horseshoe Bayou Bridge 

More than 20 CCA-treated bridges were evaluated during the 

site selection process. Most existing CCA-treated bridges are 

pedestrian in nature and were excluded for that reason. Because metal losses decline exponentially with time from 

CCA-treated wood (Brooks 1996, 1997c), it was considered 

important to evaluate at least one CCA-treated timber bridge 

Table 36—Water quality standards for surface waters 

in Washington state, based on EPA criteriaa 

Water quality (μg/L) 

Contaminant 

Fresh 

acute 

Fresh 

chronic 

Marine 

acute 

Marine 

chronic 

Arsenic 360 190 69 36 

Chromium(VI) 16.0 11.0 1,100.0 50.0 

Copper 10.4 7.3 2.5 

a

A hardness of 59 mg calcium carbonate/L was used for 

values requiring computation. 

Table 37—Marine water quality criteria and ambient 

concentrations of dissolved metals in Puget Sound 

seawatera

 

 Ambient concentrations of dissolved metals 

(μg/L) 

Metal 

EPA water 

quality criteria 

Puget Sound 

ambient levels 

Arsenic 36.0 1.50 

Copper 2.5 0.25 

Zinc 58.0 0.50 

a

U.S. EPA 1985. 

Table 38—Comparison of marine sediment quality 

criteria (AET) and benchmarks (TEL and PEL) 

Value (μg metal/g dry sediment) 

Metal AETa TELb PELb (TEL + 

PEL)/2 

Copper 390 18.7 108 63.35 

Chromium 260 52.3 160 106.15 

Arsenic 57 7.24 41.6 24.42 

a

Apparent effects threshold (AET) based marine sediment 

 quality criteria defined by Washington State. 

b

Threshold effects level (TEL) and probable effects level 

 (PEL) marine sediment quality benchmarks (Jones and 

 others 1997). 

Table 39—Freshwater sediment benchmarksa 

Metal 

Threshold 

effects level 

(TEL) 

Probable 

effects level 

(PEL) (TEL+PEL)/2 

Copper 28 100 64 

Chromium 36 120 78 

Arsenic 11 48 30 

a

Ingersol and others (1996). These are not regulatory 

 criteria. They are benchmarks for screening sediments 

 based on amphipod and midge bioassays. 
72 

Figure 36—Site plan showing the location of the Horseshoe Bayou Bridge (marine) and the Fountains 

Bridge (freshwater) located in Sandestin Resort, Sandestin, Florida. 

within a week of completing construction. An HS20 bridge, 

treated entirely with CCA and scheduled for completion in 

March 1998 was located in Sandestin, Florida, at the 

Sandestin Resort. The location of the bridge is identified in 

Figure 36 along with the second CCA-treated bridge examined in this evaluation. The Horseshoe Bayou Bridge crosses 

Choctawhatchee Bay and provides access to Jolee Island. 

The total length is 84.6 m (282 ft). A copy of the bridge 

elevation plan is provided in Figure 37. The Horseshoe 

Bayou Bridge (HBB) sits on 87 Class A southern yellow pine 

piling treated to a retention of 40 kg/m3

 (2.5 lb/ft3) with CCA 

type C. Piling cross bracing was treated to 12.8 kg/m3

 

(0.8 lb/ft3

) with CCA-C; the bridge deck, support beams, and 

railings were treated to 6.4 kg/m3

 (0.40 lb/ft3) CCA. The 

bridge was nearing completion on March 22, 1998, when 

the risk assessment evaluation was conducted. 

Rainfall in Sandestin, Florida 

February and March of 1998 were wetter than normal in 

Sandestin, Florida. Table 40 describes rainfall patterns prior 

to the time of evaluating the new Horseshoe Bayou CCAtreated bridge. The weather was clear and sunny during this 

evaluation. However, 38 mm (1.5 in.) of rain fell on February 

16, 1998, and a total of 110 mm (4.23 in.) fell during the 

Horseshoe Bayou Bridge 

Fountains Bridge 
73 

Table 40—Sandestin, Florida, rainfall 

 Rainfall (mm (in.)) 

Month Average 1998 

January 113.5 (4.47) 255.0 (10.04) 

February 124.4 (4.90) 245.1 (9.65) 

March 143.8 (5.66) 203.5 (8.01) 

preceding 2 weeks. Rainfall was documented because most 

of the structure is located above water. The fact that significant rainfall occurred during construction and immediately 

prior to evaluating this bridge is important because it exacerbates the potential for metal loss from the overhead structure. 

Figure 38 shows that portion of the bridge crossing Choctawhatchee Bay. Figure 39 is a copy of the elevation plan for 

this bridge with the location of the sampling transect annotated. Water depths in Horseshoe Bayou under the bridge 

varied between 0.0 and 104 cm. The salinity was 25.5 ppt 

and the water temperature 15.8o

C. The pH of the water was 

6.9. Currents were measured using a drogue. They were less 

than 1.0 cm/s during slack tide and increased to 2.5 cm/s 3 h 

following slack. The water depths given in Table 41 were 

measured during low slack tide. All suspended solids in the 

water were organic (volatile). 

Dissolved Copper, Chromium, and Arsenic 

Figure 39 summarizes the concentrations of dissolved metals 

in the water column as a function of distance from the Horseshoe Bayou Bridge. The U.S. EPA chronic marine water 

quality criteria are 2.5 μg copper/L, 36 μg arsenic/L, and 

50 μg chromium(VI)/L. 

All observed water column metal concentrations were close 

to background, and none approached its respective criterion. 

These results suggest that there were no adverse biological 

effects associated with dissolved metals lost from this newly 

constructed bridge treated with CCA-C. Differences in the 

concentration of dissolved metals were not significant between any two stations. Metal concentrations were slightly 

higher further from the bridge. 

Sedimented Copper, Chromium, and Arsenic 

During construction of this bridge, 1,568 holes, each 19 mm 

(0.75 in.) in diameter and 0.35 m (14 in.) long, were drilled 

in the treated wood to secure structural members with galvanized bolts. Shavings from these borings were not cleaned 

up, and they blew into the estuary, forming mats in the grass 

(Fig. 40). No effort was made to avoid these shavings during 

sampling. Sediment physicochemical characteristics observed on March 22, 1998, are summarized in Table 42. 

Sediment concentrations of copper, chromium, and arsenic 

are higher within 3 m (10 ft) of the bridge. The distribution 

Figure 37—Elevation drawing of the Horseshoe Bayou Bridge in Sandestin Florida. The sampling transect is shown. 
74 

Figure 39—Dissolved copper, chromium, and arsenic 

observed in salt water as a function of distance from 

the Horseshoe Bayou Bridge in Sandestin, Florida.

of these metals was very patchy, and high values were associated with variance to mean ratios that significantly exceed 

one. The ratio of CuO:CrO3:As2O5 in CCA is 18.5:47.4:34 

by weight (AWPA 1996). In all except results from the sample collected 0.9 m (3 ft) from the bridge, the ratio of metals 

in these sediment samples approximated that of CCA-C. The 

presence of the wood chips, coupled with the patchy distribution and the ratio of metals, suggests that the observed metals 

were in fact associated with the debris and were not the result 

of loss from the preserved wood followed by sorption and 

sedimentation. These metals remain fixed in the wood debris 

and were not generally bioavailable. However, this debris 

represents an “unnecessary environmental risk.” Sediment 

concentrations of copper, chromium, and arsenic are summarized in Figure 41. 

Sediment concentrations of metals observed in the vicinity of 

the Horseshoe Bay Bridge are summarized in Table 43 along 

with sediment benchmarks from Table 38. The maximum 

sediment concentration of copper did not exceed the TEL. 

However, the maximum concentrations of chromium and 

arsenic did exceed their respective TEL, and it was possible 

that adverse effects might be observed in the biota if this 

metal was bioavailable. However, from a realistic point of 

view, mean metal concentrations at all stations were less than 

or equal to the (TEL + PEL)/2 benchmark used in this assessment, and adverse effects were considered unlikely. This 

is especially true in light of the preceding discussion regarding the nature of the nearfield metals and the likelihood that 

they were fixed in the shavings and not bioavailable. 

Figure 38—The Horseshoe Bayou Bridge crossing 

Choctawhatchee Bay in Sandestin, Florida. The 

sampling transect is noted. 

Table 41—Physicochemical characteristics observed on March 22, 1998, in the water column at Horseshoe 

Bayou Bridge, Choctawhatchee Baya

 

Station 

Depth 

(cm) pH 

Cu 

(μg/L) 

Cr 

(μg/L) 

As 

 (μg/L) 

TSS 

(mg/L) 

TVS 

(mg/L) 

Salinity 

(ppt) (o

C) 

−52.5 m (−175 ft) control 30 8.1 2.03 + 0.35 1.80 + 0.25 8.27 + 2.10 3.6 3.6 25.0 15.8 

0.0 under bridge 40 8.1 1.55 + 0.10 1.72 + 0.35 6.19 +0.43 25.0 15.8 

0.45 m (1.5 ft) downstream 40 8.1 1.75 1.98 6.97 10.0 10.0 25.0 15.8 

0.9 m (3.0 ft) downstream 30 8.1 1.59 2.10 6.34 25.0 15.8 

1.8 m (6.0 ft) downstream 30 8.1 1.80 1.88 6.38 5.2 5.2 25.1 15.8 

3 m (10 ft) downstream 30 8.1 1.92 2.02 6.71 25.1 15.8 

6 m (20 ft) downstream 30 8.1 1.86 1.95 7.41 25.0 15.8 

9.9 m (33 ft) downstream 30 8.1 1.92 1.78 7.16 4.4 4.4 25.0 15.8 

a

The control and samples taken under the bridge are the mean of three replicates. Where appropriate, 

 values include 95% confidence limits on the mean. 

0

1

2

3

4

5

6

7

8

9

0 0.45 0.9 1.8 3.0 6.0 9.9 53.3

Distance from Horseshoe Bayou Bridge (m)

Metal concentration (

μg/L)

Copper Chrome Arsenic
75 

Figure 41—Sediment concentrations of copper, 

chromium, and arsenic observed in the vicinity 

of Horseshoe Bayou Bridge in Sandestin, Florida. 

Biological Response—Infauna

A total of 1,537 invertebrates representing 46 taxa were 

identified in the 18 samples covering 0.0309 m2

 each. Taxa 

representing a minimum of 1% of the total (15 individuals) 

were considered dominant. Table 44 describes the dominant 

taxa observed in these samples. 

Dominant taxa represent 89% of the total invertebrate abundance in these samples. The invertebrate community is dominated by polychaetes, particularly deposit feeders that might 

be sensitive to sedimented metals that are bioavailable. The 

number of taxa, total species abundance, and dominant species abundance are displayed in Figure 42 as a function of 

distance upstream and downstream from the Horseshoe 

Figure 40—CCA-treated wood shavings deposited in 

Choctawhatchee Bay as a result of in situ boring of 

the bridge structure during construction. 

Table 42—Physicochemical characteristics observed on March 22, 1998, in sediments at Horseshoe Bayou Bridge, 

Choctawhatchee Baya

 

Station 

Depth 

(cm) 

Cu 

(mg/kg) 

Cr 

(mg/kg) 

As 

(mg/kg) 

TVS 

 (%) 

Sand 

 (%) 

Silt and clay 

 (%) 

−52.5 m (−175 ft) control 30 2.43 +

0.43 

4.13 +

0.82 

0.80 +

0.11 

1.34 + 0.28 84.38 + 4.21 14.27 + 2.12 

0.0 under bridge 40 11.87 +

13.01 

23.57 +

30.16 

9.80 +

14.90 

2.95 + 1.33 79.47 + 4.63 20.19 + 4.43 

0.45 m (1.5 ft) downstream 40 7.95 +

11.60 

17.20 +

26.40 

2.65 +

3.60 

0.23 91.76 8.24 

0.9 m (3.0 ft) downstream 30 4.25 +

0.72 

10.80 +

0.64 

17.90 + 

20.96 

5.12 88.35 11.14 

1.8 m (6.0 ft) downstream 30 3.80 +

3.84 

6.93 +

7.84 

3.80 +

5.12 

1.19 87.75 12.17 

3 m (10 ft) downstream 30 1.55 +

0.72 

2.30 +

1.44 

0.85 +

0.08 

0.43 90.44 9.56 

6 m (20 ft) downstream 30 2.15 +

0.24 

3.20 +

0.48 

0.70 +

0.00 

0.81 86.33 13.30 

9.9 m (33 ft) downstream 30 1.80 +

0.16 

3.15 +

0.40 

0.55 +

0.08 

0.99 87.65 12.35 

a

The control and samples taken under the bridge are the mean of three replicates. Where appropriate, 

 values include 95% confidence limits on the mean. 

0

5

10

15

20

25

0 0.45 0.9 1.8 3.0 6.0 9.9 53.3

Distance from bridge (m) 

Metal concentration 

(μg/g dry sediment) 

Copper Chrome Arsenic 
76 

Bayou Bridge. Sediment values of arsenic were included 

because this was the only metal that exceeded its respective 

benchmark, albeit by a small amount (31/30). 

There was a shallow cline in the number of taxa observed in 

the vicinity of the Horseshoe Bayou Bridge. More taxa were 

found under the bridge than were found at further distances. 

This cline was investigated using linear regression analysis 

on log(N = 1) transformed count data; log transformed continuous data; and arcsine(square root(proportion)) transformed proportional (TVS and fines) data. The number of 

taxa was determined only by the constant term and the depth 

of the reduction–oxidation potential discontinuity in a linear 

regression. The null hypothesis that the constant term and 

coefficient on RPD were zero was rejected at α = 0.05, but 

the regression explained only 24% of the variation in the 

database. Total abundance and dominant taxa abundance 

were not significantly different (α = 0.05) between the three 

replicates collected under the bridge and those collected at 

the reference site. 

The relationships between Shannon’s and Pielou’s Indexes 

and sediment physicochemical parameters are explored in 

Figure 43. These results are consistent with the hypothesis 

that the observed metals were associated with wood fibers 

(drill shavings) from bridge construction. This statement is 

supported by the exceptionally high value of TVS at Station 

0.9 m (3 ft), where the high metal content was also observed. 

The proportion sediment silt and clay (FINES) was not elevated at this station, suggesting that the increase was not 

associated with the accumulation of detritus. The metals 

bound in the wood fibers would not be biologically available, 

and that explains the lack of significant negative correlation 

between biological endpoints and sediment metal concentrations. Shannon’s and Pielou’s Indexes were highest under the 

bridge and out to 0.9 m (3 ft) from the perimeter in the area 

where the highest concentrations of metals were also observed. This would be an unlikely finding if the metals were 

biologically available and of sufficiently high concentration 

to create adverse effects. 

Table 43—Mean and maximum concentrations of copper, 

chromium, and arsenic observed in sediments adjacent to 

the newly constructed Horseshoe Bayou Bridge in Sandestin, Florida, compared with TEL and PEL 

Metal 

Maximum 

value 

(mg/kg) 

Maximum 

mean 

(mg/kg) PEL TEL 

(TEL + 

PEL)/2 

Copper 25.1 11.87 100.00 28.00 64.00 

Chromium 54.3 23.57 120.00 36.00 78.00 

Arsenic 31.0 17.90 48.00 11.00 30.00 

Table 44—Dominant invertebrate taxa observed in sediment samples collected in the vicinity of Horseshoe 

Bayou Bridge in Sandestin, Floridaa

 

Dominant taxa Code 

Number of 

samples 

Total 

abundance Tropic group 

Arthropods 

Hargeria rapax AHARGPAX 15 51 

Grandidierella bonnieroides AGRANDIB 12 59 

Mollusks 

Neritina usnea MGNERITI 10 44 Vegetarian 

Tagelus plebeius MBTAGPLE 9 19 

Eschadium reccurvum MISCHREC 2 22 

Polychaetes 

Neanthes succinea PNEASUC 17 119 Omnivore 

Neanthes micromma PNEAMIC 17 393 Omnivore 

Capitella capitata PCAPCAP 12 38 Detritivore 

Mediomastus ambiseta PMEDAMB 15 161 Detritivore 

Ampharete americana PAMPHAME 18 369 Detritivore 

Leitoscoloplos robustus PLEITORO 9 22 Detritivore 

Leitoscoloplos fragilis PLEITOFR 15 72 Detritivore 

Abundance of dominant taxa 1,369 

Total abundance of all taxa 1,537 

a

Eighteen samples were collected, each covering 0.0309 m2. 
77 

Figure 42—The number of taxa, the abundance of all 

taxa, and only dominant taxa compared with sediment 

concentrations of arsenic observed at the Horseshoe 

Bayou CCA-C-treated bridge in Sandestin, Florida. 

Figure 43—Biological endpoints (Shannon and Pielou 

Indexes) compared with sediment arsenic, total volatile 

solids (TVS), and percentage of silt and clay (FINES) as a 

function of distance from the recently constructed 

Horseshoe Bayou Bridge in Sandestin, Florida. 

A strong acid sample digestion (EPA 3000 Series) procedure 

was used in the analysis of metals. This procedure releases 

copper, chromium, and arsenic from wood fibers. 

There was an observable decrease in the abundance of invertebrates in sediments located at intermediate distances from 

the perimeter of the bridge (stations 0.45, 0.9, 1.8 m) . To 

investigate this further, transformed data were subjected to 

correlation analysis with all biological endpoints along one 

vector and sediment physicochemical parameters along the 

other. The resulting matrix is presented in Table 45. The 

number of taxa and Shannon’s Index are negatively correlated with the depth of the RPD. This was unexpected because the deeper the RPD, the more “room” there is in the 

sediment column supporting aerobic metabolism. However, 

the minimum depth of the RPD in these sediments was 

1.0 cm, which in the author’s experience is sufficient to 

support a healthy invertebrate community. The only significant correlation between sediment metals and biological 

response was the positive correlation between the abundance 

of Capitella capitata and sediment concentrations of arsenic. 

The relationship between sediment physicochemical characteristics and biological response was summarized using the 

varimax normalized principal factors (MINRES) analysis 

(Table 46). Loadings >0.60 are in bold type. The first two 

factors explained 45.4% of the variation. The next two factors explained an additional 21.7%, giving a total of 67.2% 

for the four factors considered in this analysis. 

Factor 1 is defined by physicochemical variables measuring 

the concentration of sedimented metals and TVS in the positive direction. Sediments in this area were reasonably homogeneous. The strong correlation between sediment metal 

levels and TVS was probably associated with treated wood 

shavings lost during drilling numerous holes in the structure 

during construction. These metals would not be biologically 

available and therefore would have little effect on the infaunal community. Sediments under the bridge contained more 

fines and less oxygen (the RPD was not as deep). However, 

these loadings were not significant. 

Factor 2 is defined primarily by biological variables. The 

correlations described by these loadings suggest that the 

abundance of Neanthes micromma, Mediomastus ambiseta, 

and Ampharete americana were significant factors in determining the total taxa abundance and dominant species abundance. The arthropod Hargeria rapax and the number of taxa 

were also moderately positively correlated with the significant factors. Note that these orthogonal factors are independent and that sediment concentrations of metals (copper, 

chromium, and arsenic) have very low loadings on factor 2 

and therefore have little influence on these biological 

metrics. 

Factors 3 and 4 suggest interesting ecological relationships, 

but metals are not significantly loaded on either of these 

factors, suggesting that the relationships are independent of 

sediment concentrations of copper, chromium and arsenic, at 

least at the levels observed in this study. 

The reason for the independence of the biological variables 

expressed in factors 2, 3, and 4 from the physicochemical 

variables described in factor 1 is that the metal levels are low 

and much of the metal observed is likely fixed in the wood 

shavings and not bioavailable. 

Figure 44 displays the relationship between the variables and 

factors 1 and 2. The distance between Shannon’s Index and 

measures of abundance (LABUND and LDABUND) suggests that increases in abundance were primarily dominated 

by a few species—particularly the polychaetes Neanthes 

 

0

20

40

60

80

100

120

0 0.45 0.9 1.8 3.0 6.0 9.9 53.3

Distance from Horseshoe Bayou Bridge (m)

Number taxa or invertebrates

0

10

20

30

40

50

60

70

80

90

100

Sediment arsenic (

μg/g)

Taxa Total abundance

Dominant abundance Sediment arsenic

0

1

2

3

4

5

6

0 0.45 0.9 1.8 3.0 6.0 9.9 53.3

Distance from Horseshoe Bayou Bridge (m)

Shannon’s and Pielou’s Indexes 

and TVS (%)

0

5

10

15

20

25

Fines (%) and sedimented 

arsenic (

μg/g)

Shannon’s Index

Pielou's index

TVS

Sediment arsenic

Fines

Current direction 
78 

Table 45—Matrix of Pearson correlation coefficients comparing biological endpoints, including individual 

dominant taxa and all metrics with sediment metal concentrations, depth of the aerobic layer (RPD), total 

volatile solids (TVS), and percentage silt and clay at the Horseshoe Bayou Bridge in Sandestin, Florida 

 RPD Copper Chromium Arsenic TVS 

Percentage 

silt and clay 

Hargeria rapax −0.03 −0.03 −0.03 0.07 −0.13 −0.27 

Grandidierella bonnieroides 0.39 −0.16 −0.19 −0.37 −0.09 0.37 

Neritina usnea −0.44 0.35 0.36 0.46 0.03 −0.35 

Tagelus plebeius 0.02 0.09 0.06 −0.09 −0.18 0.10 

Ischadium reccurvum −0.42 0.12 0.17 0.41 0.35 0.06 

Neanthes succinea −0.24 0.07 0.06 0.13 −0.11 −0.34 

Neanthes micromma 0.04 −0.03 −0.04 0.02 −0.14 −0.13 

Capitella capitata −0.46 0.40 0.36 0.47 0.09 −0.01 

Mediomastus ambiseta −0.03 −0.28 −0.27 −0.10 0.15 0.03 

Ampharete americana −0.28 −0.04 −0.05 0.08 0.23 0.16 

Leitoscoloplos robustus 0.40 −0.28 −0.30 −0.37 −0.38 −0.29 

Leitoscoloplos fragilis 0.33 −0.37 −0.31 −0.26 0.02 −0.21 

Total abundance −0.06 0.07 0.07 0.09 −0.01 −0.05 

Number taxa −0.51 0.39 0.35 0.30 0.21 0.11 

Shannon’s index −0.49 0.25 0.25 0.31 0.16 −0.15 

Pielou’s index −0.17 −0.02 0.01 0.15 0.00 −0.32 

Dominant abundance 0.03 −0.03 −0.03 0.02 −0.05 −0.07 

Table 46—Summary of loadings on four factors associated with a varimax normalized principal factors 

(MINRES) analysis of physicochemical and biological variables measured in sediments adjacent to 

Horseshoe Bayou Bridge in Sandestin, Floridaa

 

 Factor 1 Factor 2 Factor 3 Factor 4 

Hargeria rapax (AHARGPAX) −0.08 0.49 0.02 0.40 

Grandidierella bonnieroides (AGRANDIB) −0.11 0.14 −0.94 0.01 

Neritina usnea (MGNERITI) 0.22 0.09 0.62 0.55 

Neanthes succinea (PHEASUC) −0.05 0.18 0.70 0.25 

Neanthes micromma (LPNEAMIC) −0.03 0.71 0.24 −0.25 

Mediomastus ambiseta (LPMEDAMB) −0.12 0.62 0.08 −0.08 

Ampharete americana (LPAMPHAME) 0.13 0.63 −0.20 0.01 

Leitoscoloplos fragilis (LPLEITOFR) −0.32 −0.01 −0.49 0.38 

Total taxa abundance (LABUND) 0.05 0.93 0.00 0.30 

Number of taxa (LDIVER) 0.36 0.47 0.10 0.62 

Shannon's Index (LSHANNON) 0.19 −0.07 0.17 0.95 

Depth of the RPD (LRPD) −0.51 −0.04 −0.28 −0.26 

Sediment copper concentration (LSEDCU) 0.87 −0.05 0.11 0.08 

Sediment chromium concentration (LSEDCR) 0.87 −0.06 0.12 0.09 

Sediment arsenic concentration (LSEDAS) 0.82 0.01 0.25 0.14 

Sediment total volatile solids (LSEDTVS) 0.66 0.01 −0.12 0.02 

Sediment silt and clay (LFINES) 0.59 0.00 −0.41 −0.25 

Dominant species abundance (LDABUND) −0.04 0.97 −0.01 0.16 

Explained variation 3.62 3.64 2.51 2.32 

Proportion of total variation 0.20 0.20 0.14 0.13 

a

Loadings in bold type are >0.60. 
79 

Figure 44—Varimax normalized principal pactors (MINRES) analysis of selected 

physicochemical and biological parameters, displaying the relationship between 

factors 1 and 2. Variable codes are given in Appendix B. 

micromma, Ampharete americana, and Mediomastus ambiseta and the arthropod Hargeria rapax. 

Laboratory Bioassays 

Water samples for Menedia berylina bioassays were collected within 15 cm of the piling at low slack tide using a 

portable peristaltic pump and precleaned, acid-washed platinized silicone tubing. Water samples were filtered in the field 

through Corning Costar Membra-Fil filters (0.45 μm). The 

bioassay results are summarized in Table 47. 

Survival was excellent in all tests, and there were no statistically significant differences in survival between any of the 

stations. These results are consistent with the low water 

column concentrations of metals observed in the Horseshoe 

Bayou water. 

Summary of Biological Response 

Marine organisms are more susceptible to copper intoxication than are most freshwater taxa. This is reflected in the 

lower U.S. EPA dissolved copper standard in salt water 

(2.5 μg/L) than in freshwater (generally 5 to 20 μg/L). Therefore, this large structure, located in an area of very slow 

currents, represents a worst-case analysis. This evaluation 

was completed during construction of the bridge when metal 

loss rates from CCA-treated wood would be highest. Only 

background concentrations of copper, chromium, and arsenic 

were observed dissolved in the water column. This suggests 

that the wood was well treated and fixed prior to installation. 

Moderately elevated levels of copper, chromium, and arsenic were observed in sediments adjacent to the bridge. The 

results of this analysis suggest the association with deposits 

of CCA-treated wood shavings lost from the bridge during 

drilling of 1,568 holes used to attach bracing and railings. 

The metals in these shavings were fixed to the wood fibers 

and likely had low bioavailability. However, this represents 

unnecessary risk and construction contracts should be conditioned to require immediate cleanup of this type of debris. 

Adverse biological effects were not observed in infauna at 

the Horseshoe Bayou Bridge. Adverse effects would most 

likely have been observed during this study, conducted during construction, when metal loss rates were the highest. 

Principal factor’s analysis indicated the biological endpoints 

Table 47—Summary results of Horseshoe Bayou water 

column bioassays using Menidia berylinea 

 Surviving amphipods (%) 

Replicate 

Laboratory 

control 

−52.5 m 

(−175 ft) 

control 

+0.45 m 

(+1.5 ft) 

+1.8 m 

(+6.0 ft) 

+9.9 m 

(+33.0 ft) 

1 90 90 80 80 90 

2 82 100 90 90 70 

3 100 80 90 90 100 

4 100 100 90 100 100 

Mean 93 92.5 87.5 90.0 90.0 

a

An initial amphipod count of 10 or 11 per replicate. 

Factor 1

Factor 2

LAHARGPA

LAGRANDI

LMGNERIT

LPHEASUC

LPNEAMIC

LPMEDAMB LPAMPHAM

LPLEITOF

LABUND

LDIVER

LSHANNON LRPD LSEDCU

LSEDCR

LSEDTVSLSEDAS LFINES

LDABUND

-0.2

0

0.2

0.4

0.6

0.8

1.0

1.2

-0.8 -0.6 -0.4 -0.2 0 0.2 0.4 0.6 0.8 1.0
80 

were independent of sediment or water column concentrations of copper, chromium, or arsenic. Shannon’s Index and 

Pielou’s Index were slightly higher within 0.9 m (3 ft) of the 

bridge compared with further distances, including the reference station. Similarly, the abundance and/or number of taxa 

observed at stations in the immediate vicinity of the bridge 

were as high as or higher than the same metrics observed at 

the reference station. Survival in laboratory bioassays was 

excellent at all stations, and statistically significant differences were not observed between any two stations. 

Despite the loss of nearly 0.28 m3

 (10 ft3) of drill shavings to 

the local environment, the Horseshoe Bayou Bridge was 

having no effect on biological resources assessed in the 

evaluation. 

Results From Fountains CCA-Treated Bridge 

The Fountains Bridge crosses a eutrophic freshwater marsh 

(Fig. 36). Figure 45 is a photograph of the bridge and local 

environment with the sampling transect indicated. Figure 46 

depicts the environment adjacent to the bridge. This bridge 

was constructed 2 years prior to the evaluation. No water 

movement was detected in the vicinity of the Fountains 

Bridge, and it was anticipated what maximum sediment 

concentrations of copper, chromium, and arsenic associated 

with worst-case CCA-treated wood projects constructed in 

aquatic environments would be observed in sediments at this 

site. 

Rainfall in Sandestin, Florida 

February and March of 1998 were wetter than normal in 

Sandestin, Florida. Table 40 describes rainfall patterns prior 

to the time of evaluating the new Horseshoe Bayou CCAtreated bridge. The weather was clear and sunny during this 

evaluation. However, 38 mm (1.5 in.) of rain fell February 

16, 1998, and a total of 110 mm (4.23 in.) fell during the 

preceding 2 weeks. Rainfall was documented because most 

of the structure is located above water. The 2-year-old Fountains Bridge was expected to lose very little metal with longterm steady state loss rates of 0.114 μg/cm2

-day for copper, 

0.04 μg/cm2

-day for arsenic, and 0.015 μg/cm2-day for 

chromium (Brooks 2000). These losses were less than 5% of 

predicted initial loss rates. Therefore, although the amount of 

rainfall recorded in the period immediately preceding this 

study was significant and greater than normal, the metals lost 

in association with the structure would likely not be detectable above background concentrations because of the age of 

the structure. 

Dissolved Copper, Chromium, and Arsenic 

Table 48 summarizes physicochemical characteristics observed in the water column adjacent to the Fountains Bridge 

during March 20−23, 1998. Figure 47 summarizes the concentrations of total metals in the water column as a function 

of distance from the Fountains Bridge. 

The marshland associated with the Fountains Bridge is eutrophic, with mats of vegetation and blue green algae. Alkalinity was low and hardness was moderate at 59.0 mg calcium carbonate/L. The low pH was not expected because of 

the amount of vegetation growing in the pond. However, 

increased metal loss would not be predicted from CCAtreated wood at the values observed (Cooper 1991). At the 

measured hardness, the U.S. EPA water quality criterion for 

copper is 6.49 μg Cu/L; chronic arsenic criterion is 190 μg 

As/L; chromium(VI) standard is 11.0 Cr(VI)/L, chromium(III) criterion is134 μg Cr(III)/L. Cooper (unpublished 

data) noted that 95% of the chromium released from CCAtreated wood in leaching studies was in the chromium(III) 

valence state. In any case, increased levels of copper, chromium, and arsenic were not observed in the vicinity of the 

bridge and the background levels are well below chronic 

criteria. 

These results suggest that there would be no adverse biological effects associated with dissolved metals lost from this 

2-year-old bridge treated with CCA-C. Differences in the 

Figure 45—Fountains Bridge in Sandestin, Florida. 

Figure 46—Marshlands where the invertebrate effects 

of the Fountains CCA-treated timber bridge were 

evaluated in Sandestin, Florida. 
81 

Figure 47—Dissolved copper, chromium, and arsenic 

observed in saltwater as a function of distance from 

the freshwater Fountains Bridge in Sandestin, Florida. 

concentration of dissolved copper and arsenic were not 

significantly different between any two stations. and maximum concentrations of arsenic were less than 20% of the 

current drinking water standard (50 μg/L). Chromium concentrations increased with distance from the bridge. Dissolved chromium was significantly higher at a distance of 

30 m (100 ft) from the bridge compared with the 0.9-m (3.0-

ft) station. Regression analysis indicated that the apparent 

slope observed in the chromium data was statistically significant. The null hypothesis that either the constant term or the 

coefficient on distance were zero was rejected at α = 0.05 

(PCr = 0.004 and Pconstant = 0.000). The resulting regression 

was 

Dissolved Chromium (μg/L) = 0.090 + 0.001 × Distance 

Sedimented Copper, Chromium, and Arsenic 

Sediment physicochemical characteristics are summarized in 

Table 49, and sediment concentrations of copper, chromium, 

and arsenic are provided in Figure 48. Sediment levels of 

copper, chromium, and arsenic were slightly elevated to a 

distance between 3 and 10 m (10 and 20 ft) from the perimeter of the Fountains Bridge. 

Sediment concentrations of metals observed in the vicinity of 

the Fountains Bridge are compared with sediment metal 

benchmarks in Table 50. All values, including the maxima 

observed in any sample, were less than the TEL, and no 

adverse biological effects could reasonably be expected. 

Biological Response

Sediments were heavily covered with vegetation. All vegetation and sediment to a depth of about 10 cm were removed 

from within a quadrat covering 0.03 m2

. A total of 11,843 

invertebrates representing 43 taxa were identified in the 18 

samples. The invertebrate community was dominated by 

annelids (7,975 total). Taxa representing a minimum of 1% 

of the total, excluding annelids (11 taxa), were considered 

subdominant. Table 51 describes the dominant and subdominant taxa observed in these samples. 

Dominant and subdominant taxa accounted for 98.2% of 

total abundance and included only 13 of the 43 taxa observed 

in 18 samples. Excluding annelids, the remaining 12 taxa 

represented 31% of the total abundance. All measures of 

diversity were based on the entire database. Abundance 

measures were analyzed using both total abundance and the 

abundance of only subdominant species. 

Table 48—Physicochemical characteristics observed in the water column at the Fountains Bridge in Sandestin, Florida 

on March 23, 1998a 

Station 

Depth 

(cm) 

Hardness 

(mg/L) 

Alkalinity 

(mg/L) pH 

Copper 

(μg/L) 

Chromium 

(μg/L) 

Arsenic 

(μg/L) 

TSS 

(mg/L) 

TVS 

(mg/L) (o

C) 

0.0 under bridge 18.3 59.0 + 21.5 17.1 6.40 2.11 0.68 5.73 1.52 1.52 15.8 

0.45 m (1.5 ft) downstream 15.0 59.0 + 21.5 17.1 6.41 2.62 0.80 7.12 1.52 1.52 15.8 

0.9 m (3.0 ft) downstream 17.5 59.0 + 21.5 17.1 6.35 2.15 0.44 5.91 1.52 1.52 15.8 

1.8 m (6.0 ft) downstream 17.5 59.0 + 21.5 17.1 6.20 2.03 1.00 5.37 1.52 1.52 15.8 

3 m (10 ft) downstream 30.0 59.0 + 21.5 17.1 6.25 2.53 1.61 6.52 1.52 1.52 15.8 

6 m (20 ft) downstream 20.0 59.0 + 21.5 17.1 6.25 2.29 1.52 6.11 1.52 1.52 15.8 

9.9 m (33 ft) downstream 20.0 59.0 + 21.5 17.1 6.29 1.40 1.37 7.48 1.52 1.52 15.8 

30 m (100 ft) reference 18.3 59.0 + 21.5 17.1 6.08 2.29 2.08 6.64 1.52 1.52 15.8 

a

The +30 m (+100 ft) reference station and samples taken under the bridge are the mean of three replicates. 

 Where appropriate, values include 95% confidence limits on the mean. 

0

1

2

3

4

5

6

7

8

0 0.45 0.9 1.8 3.0 6.0 30.5 

Distance from Fountains Bridge (m) 

Metal concentration (

μg/L) 

Copper Chrome Arsenic 
82 

Figure 48—Sediment concentrations of copper, 

chromium, and arsenic observed in the vicinity of the 

Horseshoe Bayou Bridge in Sandestin, Florida. 

Table 49—Sediment physicochemical characteristics observed as a function of distance from the CCA-C-treated 

Fountains Bridge in Sandestin, Florida 

Station 

Depth 

(cm) 

Copper 

(mg/kg) 

Chromium 

(mg/kg) 

Arsenic 

(mg/kg) TVS (%) Sand (%) 

Silt and clay 

(%) 

0.0 under bridge 18.3 2.10 + 1.08 3.23 + 1.22 1.50 + 0.74 1.13 + 0.37 88.1 + 5.4 8.3 + 0.65 

0.45 m (1.5 ft) downstream 15.0 1.65 + 0.40 2.60 + 0.80 4.30 + 5.12 1.61 91.5 9.1 

0.9 m (3.0 ft) downstream 17.5 2.20 + 0.80 2.90 + 1.12 1.45 + 0.72 1.12 92.1 7.6 

1.8 m (6.0 ft) downstream 17.5 2.10 + 1.28 2.00 + 0.48 1.25 + 0.72 1.02 92.2 7.8 

3 m (10 ft) downstream 30.0 1.40 + 0.32 1.10 + 0.16 0.45 + 0.08 0.95 87.8 12.1 

6 m (20 ft) downstream 20.0 0.70 + 0.16 1.30 + 0.48 0.40 + 0.00 1.09 89.3 10.5 

9.9 m (33 ft) downstream 20.0 1.20 + 0.32 1.05 + 0.56 0.63 + 0.28 0.85 89.0 10.8 

30 m (100 ft) reference 18.3 0.63 + 0.46 1.00 + 0.48 0.57 + 0.17 1.42 90.6 9.3 

Table 51—Dominant and subdominant invertebrate taxa 

observed in sediment samples collected in the vicinity 

of the Fountains Bridge in Sandestin, Floridaa 

Dominant and 

subdominant Taxa Code 

Number 

of 

samples 

Total 

abundance 

Nematodes NEMA 8 37 

Annelida ANNE 18 7,975 

Mollusks 

Pseudosuccinea sp. MGPSEU 11 54 

Gyraulus sp. MGGYS 15 779 

Insecta 

Ephemeroptera 

Caenis sp. AICAE 15 103 

Diptera (Chironomoidea) 

Bezzia sp. AICBEZ 18 154 

Family Chironomidae AICHR 16 152 

Larsia sp. AILRS 17 182 

Paramerina sp. AIPAAS 15 151 

Macropelopia sp. AIMACS 15 140 

Chironomus sp. AICHSS 18 1,426 

Paratendipes sp. AICPTS 18 395 

Zavrelimyia sp. AIZAV 17 85 

Abundance of dominant and subdominant taxa 11,633 

Abundance of subdominant taxa 3,658 

Total abundance of all taxa 11,843 

a

A total of 18 samples were collected, each covering 

 0.030 m2

. 

Table 50—Mean and maximum concentrations of 

copper, chromium, and arsenic observed in sediments 

adjacent to the newly constructed Horseshoe Bayou 

Bridge in Sandestin, Florida, compared with TEL 

and PEL 

Metal 

Maximum 

value 

(mg/kg) 

Maximum 

mean 

(mg/kg) PEL TEL 

TEL + 

PEL)/2 

Copper 3.00 2.2 + 1.28 100.00 28.00 64.00 

Chromium 3.70 3.23 + 1.22 120.00 36.00 78.00 

Arsenic 7.50 4.30 + 5.12 48.00 11.00 30.00 

0 

1 

2 

3 

4 

5 

0 0.45 0.9 1.8 3.0 6.0 30.5 

Distance from Fountains Bridge (m) 

Metal concentration 

(μg/g dry sediment) 

Copper 

Chrome 

Arsenic 
83 

Figure 49—The number of taxa, the abundance of all 

taxa, and only dominant taxa compared with sediment 

concentrations of arsenic observed at the Horseshoe 

Bayou CCA-treated bridge in Sandestin, Florida. 

The number of taxa, total species abundance, and dominant 

species abundance are displayed in Figure 49 as a function of 

distance from the Fountains Bridge. There is an anomalous 

increase in all metrics at the 6-m (20-ft) station that does not 

correlate with any of the physicochemical variables. In addition, there is an apparent cline in all three metrics with decreasing values at increasing distances. Multiple regression 

indicated that none of these log transformed biological endpoints was significantly (α = 0.05) a function of distance (or 

any other independent physicochemical variables measured 

in this study). 

Values of Shannon’s and Pielou’s Indexes are compared with 

observed concentrations of dissolved arsenic in Figure 50. 

Dissolved arsenic was chosen because all other significant 

metal endpoints were positively correlated with biological 

endpoints. The log transformed abundance of the chironomids Macropelopia sp. (p = −0.60) and Larsia sp. (p = 

−0.68) were negatively correlated with dissolved arsenic. 

Statistically significant clines were not observed in arsenic 

levels in this marsh, and high values in all biological endpoints were observed at the 0.45-m (1.5-ft) station, where 

arsenic levels were slightly elevated. 

Figure 50—Biological endpoints (Shannon and Pielou 

Indexes) compared with sediment arsenic, total volatile 

solids (TVS), and percentage of silt and clay (FINES) as a 

function of distance from the recently constructed 

Horseshoe Bayou Bridge in Sandestin, Florida. 

The results of a two-dimensional correlation analysis are 

provided in Table 52. All significant correlations (α = 0.05) 

between biological endpoints and sediment metal levels and 

dissolved copper or chromium were positive, indicating 

higher values of biological endpoints associated with increased metal concentrations. As previously noted, both 

dissolved and sedimented metal levels were low in the vicinity of the Fountains Bridge. It is likely that the biological 

differences observed were due to the stochastic nature of 

sampling patchy environments or covarying biological or 

physicochemical factors not evaluated in this study. 

Table 52 suggests that there were reduced numbers of some 

taxa associated with increased fines (silt and clay), dissolved 

arsenic, and distance from the bridge. Several taxa were 

significantly more abundant at distances where sediments 

contained increased concentrations of copper and arsenic. 

These data were further explored using principle components 

analysis and principle factors (MINRES) analysis with varimax rotation. These analyses did not provide meaningful 

insight to this database. Physicochemical variables were not 

Table 52—Matrix of Pearson correlation coefficients comparing biological endpoints, including individual dominant taxa and 

all metrics with sediment metal concentrations, depth of the aerobic layer (RPD), total volatile solids (TVS), and percentage 

silt and clay (% FINES) at Fountains Bridge in Sandestin, Florida 

 AICHSS AIMACS ANNE MGGYS AICAE AICBEZ AICHR AILRS ABUND DABUND DIVERS SHANNON PIELOU 

FINES −0.595 −0.532 −0.336 −0.422 −0.744 0.101 −0.413 −0.752 −0.288 −0.263 −0.332 −0.445 −0.397 

Distance arsenic −0.440 −0.597 −0.261 −0.447 −0.364 0.135 −0.378 −0.683 −0.374 −0.505 −0.298 −0.618 −0.614 

Distance chromium −0.094 0.018 −0.277 −0.134 −0.024 −0.303 −0.073 0.088 −0.269 −0.203 −0.128 0.031 0.086 

Distance −0.111 −0.454 −0.595 −0.562 −0.484 −0.470 −0.487 −0.232 −0.535 −0.353 −0.262 −0.163 −0.114 

Sediment copper 0.108 0.321 0.507 0.327 0.283 0.425 0.236 0.142 0.412 0.245 0.005 0.007 0.005 

Sediment arsenic 0.348 0.471 0.549 0.379 0.473 0.322 0.221 0.305 0.422 0.244 0.158 0.101 0.068 

0 

200 

400 

600 

800 

1000 

1200 

0 0.45 0.9 1.8 3.0 6.0 9.9 30.5 

Distance from Fountains Bridge (m)

Number of organisms

0 

5 

10 

15 

20 

25 

Number of taxa

Abundance 

Subdominant abundance 

Number taxa 

 

0

0.20

0.40

0.60

0.80

1.00

1.20

1.40

1.60

Distance from Fountains Bridge (m)

Shannon's and Pielou's Indexes

0

1 

2 

3 

4

5 

6 

7

8

Dissolved arsenic (

μg/L)

Shannon’s Index

Pielou’s Index

Dissolved arsenic

0 0.45 0.9 1.8 3.0 6.0 9.9 30.5
84 

significantly differentiated from biological variables, consistent with the lack of clear relationships observed in the previous analysis. 

Bioassay Results 

Bioassays were completed on water samples using the cladoceran Daphnia magna. Results are given in Table 53. The 

proportions of surviving test animals were arcsine(square 

root (proportion surviving)) transformed and t-tests were 

conducted to compare survival between treatment stations 

located at +0.45, +1.8, and +22.8 m (+1.5, +6.0, and +33.0 

ft) from the bridge with the laboratory control and the local 

reference station located 30 m (100 ft) upstream from the 

bridge. The results are provided in Table 54. 

Significant differences in cladoceran survival were not 

observed between any treatment station and either 

the laboratory or the local control. These results are 

consistent with the low levels of metals found in this marsh 

and the lack of effects observed in the invertebrate community assessment. 

Summary of Physicochemical 

and Biological Evaluation 

This CCA-treated timber bridge was constructed 2 years 

prior to the evaluation. It is a substantial bridge with a span 

of 30 m (100 ft). The Fountains Bridge crosses a freshwater 

marsh with no observable water movement, at least not during the 3 days of this evaluation. This bridge was chosen 

because it represents a worst case with respect to sedimented 

copper, chromium, and arsenic. 

Water column concentrations of copper, chromium, and 

arsenic were below U.S. EPA water quality criteria, and there 

was no gradient indicating measurable loss from the bridge in 

the immediate past. Historical metal loss from the bridge was 

recorded in the sediments, with elevated concentrations of all 

three metals observed immediately adjacent to the structure 

and a clearly defined gradient that reached apparent background concentrations at some distance greater than 1.8 m 

(6 ft) but less than 3 m (10 ft) from the bridge perimeter. 

Maximum observed sediment levels for all three metals were 

low and less than their respective TELs. 

Adverse biological effects were not expected in association 

with the observed metal levels and not one was observed. 

More taxa, in higher abundance, were found closer to the 

bridge than further away. However, the regression coefficient 

on distance was not statistically significant. Statistically 

significant differences in the Daphnia magna laboratory 

Table 53—Summary results of Fountains Bridge water 

column 96-h bioassays using the test cladoceran 

Daphnia magnaa 

 Surviving animals (%) 

Replicate 

Laboratory 

control 

Upstream 

control 

+0.45 m 

(+1.5 ft) 

+1.8 m 

(+6.0 ft) 

+9.9 m 

(+33.0 ft) 

1 100 100 80 90 90 

2 90 80 80 70 100 

3 100 60 70 60 90 

4 70 100 100 100 70 

Mean 90.0 85.0 82.5 77.5 87.5 

a

Surviving animals with an initial count of 10 per replicate.

Table 54—Results of t-tests comparing survival of Daphnia magna at treatment stations located at distances of 0.45, 1.8, and 9.9 m (1.5, 6.0, and 33.0 ft) from the perimeter of the Fountains Bridge and laboratory (LAB) or local reference station (LB100) 

 Mean 

Standard 

deviation N Difference 

Standard 

deviation 

difference T df p 

LBLAB 1.35 0.28 

LB 0.45 m (1.5 ft) 1.19 0.26 4 0.150 0.52 0.58 3.00 0.60 

LBLAB 1.35 0.28 

LB 1.8 m (6 ft) 1.17 0.31 4 0.170 0.53 0.64 3.00 0.57 

LBLAB 1.35 0.28 

LB 9.9 m (33 ft) 1.27 0.24 4 0.080 0.31 0.52 3.00 0.64 

LB100 1.28 0.34 

LB 0.45 m (1.5 ft) 1.19 0.26 4 0.090 0.25 0.71 3.00 0.53 

LB100 1.28 0.34 

LB 1.8 m (6.0 ft) 1.17 0.31 4 0.109 0.15 1.44 3.00 0.24 

LB100 1.28 0.34 

LB 9.9 m (33 ft) 1.26 0.24 4 0.019 0.51 0.07 3.00 0.95 
85 

bioassay were not observed at any distance from this bridge. 

In short, the bridge has created small increases in nearfield 

sediment concentrations of copper, chromium, and arsenic 

and no biological effect on the invertebrate community located in this marsh. 

Conclusions 

Physicochemical and biological endpoints were evaluated at 

two creosote-treated bridges in Indiana, two penta-treated 

bridges in Oregon and Washington, and two CCA-treated 

bridges in Florida. These bridges were selected to present, as 

much as possible, worst-case projects with respect to preservative contamination of the water column and sediments. 

Preservative was lost from each bridge and could be detected 

at generally low levels in either the water column or in sediments. 

Polycyclic aromatic hydrocarbons (PAHs) were detected in 

sediments immediately downstream from the two creosotetreated bridges. Higher sediment concentrations were observed adjacent to the new bridge (range = 1.24 to 5.31 μg 

TPAH/g) than were observed at the 8-year-old bridge (0.98 

to 2.56 μg TPAH/g). Sediment levels of PAH exceeded the 

threshold effects level (TEL) for some compounds, and the 

concentration of phenanthrene was essentially equal to the 

sediment benchmark invoked for this assessment ((TEL + 

PEL)/2) at one station downstream from the newer bridge. 

Pipe Creek is located in an agricultural area and carries a 

heavy sediment load. The creek is shallow and slow moving, 

at least during the summer and late fall. The invertebrate 

community is a robust one dominated by annelids and chironomids. Adverse biological effects were not documented in 

the invertebrate community assessment or in laboratory 

bioassays on sediments from this bridge. The lack of biological response was attributed to the robust invertebrate community inhabiting the naturally stressful environment and to 

the low number of compounds that marginally exceeded the 

sediment benchmarks. Total PAH concentrations did not 

exceed appropriate benchmarks at either of these bridges. 

Penta was not observed in the water column at either the 

Dairy Creek or West Fork Satsop River Bridges. Very low 

levels, less than the penta TEL, were observed in sediments 

under both bridges. No adverse biological effects were 

documented at either of these bridges, even though the 

healthy invertebrate communities included large numbers 

of pollutant-sensitive taxa in the orders Ephemeroptera, 

Plecoptera, and Trichoptera. 

The CCA-treated Horseshoe Bayou Bridge is a substantial 

HS20 bridge that crosses a sensitive marine environment. 

This bridge was chosen because of the volume of treated 

wood used and because marine organisms are very susceptible to copper intoxication. This susceptibility is seen in the 

low (2.5 μg Cu/L) marine water quality standard. The 

evaluation was timed to coincide with the period during 

construction in which maximum metal losses from the treated 

wood were anticipated. Marine water in the vicinity of the 

nearly complete bridge did not contain elevated levels of 

copper, chromium, or arsenic. 

Elevated levels of copper, chromium, and arsenic were observed in association with sawdust and shavings lost from the 

Horseshoe Bayou Bridge during construction. The metals in 

these shavings were fixed in the wood fibers and not 

bioavailable. Adverse effects were not observed in either the 

indigenous infaunal community or in laboratory bioassays at 

this bridge. 

The Fountains Bridge was 2 years old at the time of this 

evaluation. It crosses a eutrophic marsh with no observable 

water movement. It was anticipated that this bridge would 

represent a worst-case analysis with respect to sediment 

concentrations of metals lost from CCA-treated wood. As 

expected with a 2-year-old structure, increased levels of 

dissolved copper, chromium, and arsenic were not observed 

in the vicinity of this bridge. Increased concentrations of 

sediment copper, chromium, and arsenic were observed 

under this bridge. Metal levels declined with distance from 

the bridge and reached background concentrations between 

1.8 and 3 m (6 and 10 ft) from the bridge’s perimeter. Maximum metal concentrations were all well below their respective TELs, and adverse effects were not anticipated. The 

biological evaluation at this bridge did not reveal any adverse 

effects. 

This study was designed to assess the environmental risks 

associated with timber bridges selected to represent worstcase conditions. In each case, preservative was lost from the 

bridges and could be detected in sediments but not in the 

water column. Sediment levels of contaminants were moderately high for the newer creosote-treated bridge and below 

TELs for all the other bridges. No adverse biological effects 

were observed in either the invertebrate community or in 

laboratory bioassays at any of these bridges. 

These results suggest that there are minimal environmental 

risks associated with preservatives lost from timber bridges. 

These bridges were located in rural areas and carried few 

vehicles per day. Therefore, other sources of pollution, discussed in the introduction of this report, were thought to be 

minimal. 

There is evidence that creosote can be lost from overhead 

bridge structures exposed to the sun during lengthy periods 

of hot weather. In addition, evidence of the deposition of 

coal tar in streams during road maintenance was obtained. 

Figure 51 depicts these losses at an unnamed bridge in 

Oregon. The lines of coal tar resulted from pouring tar onto 

the surface of cracks in the asphalt-covered roadway. Recall 

also the shavings observed under the newly constructed 

Horseshoe Bayou Bridge and the creosote oil dripping from 

the Narragansett Bay Bridge described in the introduction. 
86 

All these problems represent an unnecessary environmental 

risk. These are risks that can be minimized, or even eliminated, through better construction and maintenance practices. 

The conclusion to be reached is that timber bridges present 

little environmental risk but that best management practices

for construction and maintenance should be developed and 

widely disseminated. These best management practices 

should apply to all bridges, regardless of the construction 

materials used. 

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Appendix A—Test America 

Reporting Limits for Polycyclic 

Aromatic Hydrocarbons (PAHs) 

PNA 8310 Nonaqueous 

Method 

reporting limit 

(mg/kg) 

Special 

reporting limit 

(mg/kg) 

Naphthalene 0.025 0.020 

Acenaphthene 0.660 0.020 

Acenaphthylene 0.660 0.020 

Anthracene 0.660 0.020 

Phenanthrene 0.660 0.020 

Pyrene 0.180 0.020 

Benzo(a)anthracene 0.0026 0.0026 

Benzo(b)fluoranthene 0.0036 0.0036 

Benzo(k)fluoranthene 0.0034 0.0034 

Benzo(a)pyrene 0.0046 0.0046 

Benzo(ghi)perylene 0.051 0.020 

Chrysene 0.030 0.020 

Dibenzo(a,h)anthracene 0.006 0.006 

Fluoranthene 0.660 0.020 

Fluorene 0.140 0.020 

Ideno(1.2.3-cd)pyrene 0.0086 0.0086 
97 

Appendix B—Freshwater Taxonomy Codes 

CHYD Class HYDROZOA AISTEN Stenacron species 

NEMA Phylum NEMATODA AIBAET Family BAETIDAE 

ANNE Phylum ANNELIDA AIBAE cf. Baetis species 

MGJGS Juga species AIBAESS Baetis sp. 

MGLTS Leptotoxis species AICAL Callibaetis species 

MGPLS Pleurocera species AIAME Ameletus species 

MGSTL Order STYLOMMATOPHORA AICOE Family COENAGRIONIDAE 

MGFRS Ferrisia species AICOES Coenagrion species 

MGLXS Lanx species AIAS cf. Aeshna species 

MGLYS cf. Lymnaea species AIGOMS Gomphus species 

MGSTS Stagnicola species AILS Libellula species 

MGPHS Physa species AIPLE Order PLECOPTERA 

MGPHLS Physella species AINEM Family NEMOURIDAE 

MGGYS Gyraulus species AIAMS Amphinemura species 

MGHEL cf. Helisoma species AIOS Ostrocerca species 

MPPIS Pisidium species AIZAPS Zapada species 

MPSPH Sphaerium species AICAP cf. Capnia species 

AACAR Order ACARINA AIPEL Family PELTOPERLIDAE 

AARAN Order ARANEAE AICHL Family CHLOROPERLIDAE 

ACRUST Subphylum CRUSTACEA AISWE Sweltsa species 

ACDAPH Family DAPHNIIDAE AIPER Family PERLODIDAE 

ACDS Daphnia Species AIISOS Isoperla species 

ACCYC Suborder CYCLOPOIDA AICALC Calineuria californica 

ACAMP Order AMPHIPODA AIHESS Hesperoperla species 

ACHYS Hyalella species AICAED Family CAECILIUSIDAE 

ACCS Caecidotea species AIAEOL Family AEOLOTHRIPIDAE 

ACCAM Family CAMBARIDAE AIHET Suborder HETEROPTERA 

ACPAS Pacifastacus species AICOR Family CORIXIDAE 

ACOST Class OSTRACODA AISS Sigara species 

ADIPL Class DIPLOPODA AIGS Gerris species 

AIPA Podura aquatica Linnaeus, 1758 AIMS Microvelia species 

AINEA Neanura species AITING Family TINGIDAE [Terrestrial] 

AIISO Family ISOTOMIDAE AIMIR Family MIRIDAE [Terrestrial] 

AIENT Family ENTOMOBRYIDAE AIHOM Suborder HOMOPTERA 

AIDIC Dicyrtoma species AICIC Family CICADELLIDAE 

AICAE Caenis species AIMEMB Family MEMBRACIDAE 

AIEPHS Ephemerella species AIFLAT Family FLATIDAE 

AISERS Serratella species AIPSY Family PSYLLIDAE 

AIDRD Drunella doddsi AIAPH Family APHIDIDAE 

AIEPHA Ephemera species AIORTD Family ORTHEZIIDAE 

AIPAR Paraleptophlebia species AIMEG Order MEGALOPTERA 

AIHEP Family HEPTAGENIIDAE AISSL Sialis species (larva) 

AICYN Cynigma species AISSA Sialis species (adult fragment) 

AIEPES Epeorus species AINEU Order NEUROPTERA 
98 

Appendix B—Freshwater Taxonomy Codes—con. 

AICHEPS cf. Heptagenia species AIHEM Family HEMEROBIIDAE [Terrestrial] 

AIIRONS Ironodes species AICHRD Family CHRYSOPIDAE [Terrestrial] 

AIRHIS Rhithrogena species AICOL Order COLEOPTERA 

AIHSL Haliplus species (larva) AIHYHS cf. Hydropsyche species 

AIPELS Peltodytes species AIBS Banksiola species 

AIHYDR Subfamily HYDROPORINAE AIBRAD Family BRACHYCENTRIDAE 

AINEO Neobidessus species (adult) AIMIS Micrasema species 

AIHYSA Hydroporus species (adult) AILPS Lepidostoma species 

AIORES Oreodytes species (larva) AILIM Family LIMNEPHILIDAE 

AILSA Laccomis species (adult) AICHYS cf. Chyranda species 

AICOLY Subfamily COLYMBETINAE AIHAS Halesochila species 

AIASL Agabus species (larva) AIHYXS Hydatophylax species 

AIASA Agabus species (adult) AILES Lenarchus species 

AIRS Rhantus species AILIS cf. Limnephilus species 

AIGSL Gyrinus species (larva) AIPSS Psychoglypha species 

AIHELS Helophorus species AIUENO Family UENOIDAE 

AIHYDL Family HYDROPHILIDAE AILEP Order LEPIDOPTERA 

AIBERS Berosus species AIGEO Family GEOMETRIDAE 

AITROPS Tropisternus species AIDIP Order DIPTERA 

AIES Enochrus species AITIP Family TIPULIDAE 

AIPASS Paracymus species AILIMO Subfamily LIMONIINAE 

AISTA Family STAPHYLINIDAE AIANTS Antocha species 

AIPTI Family PTILIIDAE AICRLS Cryptolabis species 

AIECT cf. Ectopria species AIDIS Dicranota species 

AICYPHS Cyphon species AIERS cf. Erioptera species 

AIELM Family ELMIDAE AIHES cf. Hesperoconopa species 

AILARA Lara species AIPEOS Pedicia species 

AIDUBS Dubiraphia species AIRHXS cf. Rhabdomastix species 

AIHETS Heterlimnius species AITPS cf. Tipula species (adult) 

AINS Narpus species AITIPS Tipula species 

AIOPT Optioservus species ACEC Family CECIDOMYIIDAE 

AISTNS Stenelmis species ASCI Family SCIARIDAE 

AIBUP Family BUPRESTIDAE [Terrestrial] APSY Family PSYCHODIDAE 

AICOC Family COCCINELLIDAE [Terrestrial] AMS Maruina species (larva) 

AILATD Family LATHRIDIIDAE ACPERS cf. Pericoma species 

AICHMD Family CHRYSOMELIDAE APSYS Psychoda species 

AIALTN Subfamily ALTICINAE AANI Family ANISOPODIDAE 

AIDS cf. Donacia species AIBTS Bittacomorpha species 

AICUR Family CURCULIONIDAE AIPTY Ptychoptera species 

AISCO Family SCOLYTIDAE [Terrestrial] AIDSA Dixella species (adult) 

AITRI Order TRICHOPTERA AIDSL Dixella species (larva) 

AIRHAS Rhyacophila species AICRT Subfamily CERATOPOGONINAE 

AIGLO Family GLOSSOSOMATIDAE AICBEZ cf. Bezzia species 

AIGLOS Glossosoma species AIPBEZ Probezzia species 
99 

Appendix B—Freshwater Taxonomy Codes—con. 

AIHYDP Family HYDROPTILIDAE AIFOR Subfamily FORCIPOMYIINAE 

AIHYDS Hydroptila species (larva) AIATS Atrichopogon species 

AIOSL Oxyethira species (larva) ASIM Family SIMULIIDAE 

AIPOL Family POLYCENTROPODIDAE ASS Simulium species (larva OR pupa) 

AIPOSS Polycentropus species AICHR Family CHIRONOMIDAE 

AIHYL Family HYDROPSYCHIDAE AITNY Subfamily TANYPODINAE 

AIGUTS Gittipelopia species ACHS cf. Chrysops species 

AILRS Larsia species AGLS cf. Glutops species 

AIPAAS Paramerina species ASTRA Family STRATIOMYIDAE 

AICTHS cf. Thienemannimyia species ASTS Stratiomys species (larva) 

AIPRS Procladius species AEMP Family EMPIDIDAE 

AIPSS Psectrotanypus species ADOLI Family DOLICHOPODIDAE 

AICHSS Chironomus species ALONS Lonchoptera species 

AIDTS Dicrotendipes species ASYRD Family SYRPHIDAE 

AIGLS Glyptotendipes species AEPHY Family EPHYDRIDAE 

AIPPS Polypedilum species AMUS Family MUSCIDAE 

AITBS Tribelos species ATAC Family TACHINIDAE 

AIMPS Micropsectra species AHYM Order HYMENOPTERA 

AISTES Stempellinella species ABRAC Family BRACONIDAE 

AITYS Tanytarsus species AICH Family ICHNEUMONIDAE 

AIORTN Subfamily ORTHOCLADIINAE AICHLA Superfamily CHALCIDOIDEA 

AEUKS Eukiefferiella species AIENC Family ENCYRTIDAE [Terrestrial] 

APCLS cf. Parorthocladius species AIPLAT Family PLATYGASTRIDAE [Terrestrial] 

APSTS Psectrocladius species AFORM Family FORMICIDAE [Terrestrial] 

ARHS Rheocricotopus species AHAL Family HALICTIDAE [Terrestrial] 

ARHSS Rheosmittia species CPIS Superclass PISCES 

ASYNS Synorthocladius species CAUR Order URODELA 

APRS cf. Parochlus species CAAN Order ANURA 
100 

Appendix C—Marine Taxonomy Codes 

Arthropods Arthrop Polychaetes 

Hargeria rapax Ahargpax Neanthes succinea Pneasuc 

Rithropanopeus harrisii Arighhar Neanthes micromma Pneamic 

Edotea montosa Aedotmon Stremblospio benedicti Pstrembe 

Gammarus mucronatus Agammucr Capitella capitata Pcapcap 

Corophium louisianum Acoroloui Mediomastus ambiseta Pmedamb 

Melita sp. Amelita Polydora socialis Ppolysoc 

Grandidierella bonnieroides Agrandib Ampharete americana Pamphame 

Palaemonetes pugio Apalaepu Eteone heteropoda Peteohet 

Callinectes cf. similis Acallisi Leitoscoloplos robustus Pleitoro 

Unidentifiable arthropod Aunident Leitoscoloplos fragilis Pleitofr 

Chironomid larvae Achirono Parandalia americana Pparanam 

Erichsonella attenuata Aerichat Unidentifiable polychaete Punident 

Meilita cf. nitida Amelinit Diopatra cuprrea Pdiopcup 

Mollusks Pecternaria gouldi Ppecteng 

Gastropods Owenia sp. Powenia 

Neritina usnea MGNeriti Megalona sp. Pmegalon 

Vitrinella sp. MGVitrin Aricidea fragilis Paricfrag 

Texadina sphinetostoma MGTexadi Others 

Unidentifiable gastropod Mguniden Nemertea Nemertea 

Bivalvia Platyhelminthes Platyhel 

Tellina probina MBTelpro 

Tagelus Plebeius MBTagple 

Amygdalum papyrium MBAmypap 

Unidentifiable bivalve Mbuniden 

Crassostrea virginica Mbcravir 

Mulinia lateralis Mmulinla 

Ischadium reccurvum Mischrec 

Geukensia demissa Mgeukdem 

Telina versicolor Mtelinve