Chemical Contaminants Associated with Piers, Docks and Bulkheads

1. Hydrocarbons within Lake Washington

Direct causal links between bulkheads, piers and other artificial shorezone structures and hydrocarbon inputs are few. Those links that have been identified include the use of creosote-treated lumber (hydrocarbon flux) and localized increases in watercraft powered by internal combustion engines (hydrocarbon spillage and exhaust). Within urbanized regions, such as the Lake Washington watershed, the quantity of aliphatic and cyclic hydrocarbons entering the lake by boating activity and creosote-treated wood is most likely insignificant relative to the quantity of anthropogenic and biogenic hydrocarbons of all species entering the lake through urban runoff, atmospheric particulate fallout, and fluvial inputs (Latimer and Quinn 1998; Wakeham 1977; Jones et al. 1980; Green and Trett 1989). However, hydrocarbon input from boats and treated wood should not be considered trivial.

A. Creosote and PAHs - Creosote is primarily composed of polycyclic aromatic hydrocarbons (PAHs) (Hyötyläinen and Oikari 1998), which are known carcinogens, mutagens and teratogens (Chen et al. 1997; Hussain et al. 1998; Green and Trett 1989). PAHs are a common pyrolytic byproduct of all internal combustion engines and are now commonly found in most aquatic systems near industrialized and urbanized centers (Green and Trett 1989). PAHs are known to bio-accumulate proportionally to the number of carbon rings composing the parent compound, and have bio-accumulation factors ranging between 10,000 to 130,000 (Green and Trett 1989). Although PAHs are more commonly found as a contaminant within urban runoff (Latimer and Quinn 1998), boat exhaust (Smith et al. 1987), fuel spills, and other varied sources, under rare circumstances large quantities of creosote-treated lumber within confined bodies of water (such as San Diego Bay, California) can emit more PAHs (metric tons per year) than all other sources combined (Katz et al. 1995). In British Columbia, Goyette and Brooks (1998) found that significant PAH sediment contamination occurred within 7.5 meters of newly installed creosote piles in a marine inlet, but that significant biological effects on the infaunal community occurred within 0.65 meter. In the same study, Goyette and Brooks (1998) predicted that maximum sediment PAH concentrations should occur approximately 1,000 days post installation.

Overall, relatively little is known about the impacts of PAHs to aquatic organisms. Evidence for immunosuppression resulting from exposure to PAHs was reported by Arkoosh et al. (1998), who determined that chinook smolts from urban estuaries exhibited a higher cumulative mortality after exposure to the marine pathogen Vibrio anguillarum than smolts from a non-urban estuary. Tissue examinations of the chinook smolts indicated that those from the urban estuary had been exposed to higher levels of PAHs and PCBs than smolts from the non-urban estuary (Arkoosh et al. 1998). Studies of impacts to freshwater aquatic organisms by PAHs report inhibition of phytoplankton electron transport (Marwood et al. 1999); depression of growth and reproduction of Daphnia (Geiger and Buikema 1982); liver stress and formation of liver tumors and cataracts in rainbow trout (Hyötyläinen and Oikari 1998; Black et al. 1988; Laycock et al. 1999); and acute toxicity to the duckweed Lemna gibba (Ren et al. 1993). One study relating PAHs in freshwater sediments to human health quantified a greater than negligible lifetime increase in risk of skin cancer to persons swimming in areas with PAH-contaminated sediments (Hussain et al. 1998).

Although a few of these studies standardize and quantify lethal and sub-lethal effects (96 hr EC-50/LC-50 values), many studies employ unique exposure regimes and relative toxicity tests of unknown PAH mixtures, (I think a word may be missing), making comparisons between studies difficult (Malins 1982). Un-saturated PAHs appear to be most toxic, with effect concentration values for various freshwater and marine organisms ranging from 2 to 1,000 ppm, although many fishes have 96 hour LC-50 values between 10 and 100 ppm (Green and Trett 1989). Several studies have identified increased toxicity of PAHs exposed to ultra-violet light (Marwood et al. 1999; Laycock et al. 1999; Ren et al. 1993), an environmental condition more likely to occur within shallow littoral habitats. Additionally, photo-degraded hydrocarbons also tend to be more soluble than parent compounds, thus increasing their bioavailability (Green and Trett 1989).

B. Watercraft Exhaust Emissions - Little study has been given to the role of watercraft exhaust, as a source of environmental aliphatic and cyclic hydrocarbon contamination. In the late 1960s and early '70s several papers were published that discussed two-cycle outboard motor exhaust, although these documents provided only qualitative information regarding the extent of hydrocarbon input (Jones et al. 1980) (Citations A, B, C, none of which were acquired within the constraints of this effort). There seems to be agreement within this group of papers that two-cycle engine oil is a major source of hydrocarbon pollution within freshwater lakes (Jones et al. 1980), although other studies have estimated that relative to the overall hydrocarbon input into urbanized lakes such as Lake Washington, outboard engine operation likely contributes a very small fraction of total input, less than 1 percent (Wakeham 1977).Two recent studies investigating the effects of watercraft exhaust in relatively pristine water bodies (one marine and one freshwater) have been conducted. The first study detected low levels of PAHs in sediments near power boat moorings at Green Island in the Great Barrier Reef (marine), but PAHs were not detected in measurable concentrations in water and clam tissue at the same locations or within sediments away from power boat moorings (Smith et al. 1987). The second study implicated watercraft exhaust as the cause of seasonal increases in the water concentration of methyl tert-butyl ether (MTBE) at Donner Lake in California (Reuter et al. 1998), which is added to fuel mixtures to decrease emission of unburned hydrocarbons. MTBE is a known carcinogen, although very few studies have been conducted to determine this compound's effects on aquatic organisms. Note: a significant reference on this topic (Correll 1999) was located at press time, but was not incorporated into this report due to lack of time.2. Heavy Metal Contamination by CCA-Treated Wood

As a preventative to decay, lumber used to construct piers, docks, bulkheads and other structures experiencing regular marine and freshwater inundation are pressure treated with chromated copper arsenate (CCA). CCA is a mixture of metal oxides (chromium, copper and arsenic), each of which are highly toxic to marine and freshwater organisms in dissolved ionic form (Weis et al. 1998). Although the processes of pressure fixation of CCA to wood fibers is intended to prevent dissolution of toxic metals into the surrounding environment, contamination of water, sediment and biological organisms in proximity to CCA-treated wooden structures, especially within the first three weeks after installation, is common (Brooks 1994; Weis et al. 1998).

The toxicity of CCA leachates to freshwater and marine organisms is high, especially for copper ions which are toxic to most aquatic organisms even in comparatively low concentration, 10 to 100 ppb (Brooks 1996). Although the majority of toxicological research assessing the effects of copper, chromium and arsenic on aquatic biota has been conducted within marine waters, all three metals are known to be toxic to freshwater organisms (copper and arsenic slightly less toxic, chromium slightly more toxic) (Brooks 1996).

While substantial toxicity data for individual aquatic organisms/individual metal contaminants EC-50/LC-50 values exists, studies quantifying the impacts of leached metals in proximity to CCA-treated piers and bulkheads via in situ observations are few and limited to marine systems. Those studies that do exist have clearly demonstrated the capacity for these metals to leach into the environment in proximity to structures utilizing treated lumber. In certain circumstances, accumulations of metals leached from treated lumber have been detected in tissues of resident biota, although measurable impacts to individual organisms have been limited (Brooks 1996; Weis and Weis 1993). While new installations of CCA-treated wood structures can cause state and federal water quality standards for copper, chromium and arsenic to be exceeded in the short term (first 3 weeks), long term concentrations of these metals leaching from treated lumber is expected to remain below the effect levels for most aquatic organisms, especially in aquatic environments that are well flushed (Brooks 1996).

Three potentially useful literature reviews of the environmental risks associated with three common wood-preservatives used in aquatic environments were identified following the substantial completion of the present review. Brooks (1995a) reviewed creosote-treated wood products, Brooks (1995b) reviewed CCA-treated wood products, and Brooks (1995c) reviewed ammoniacal copper zinc arsenate (ACZA)-treated wood products. These review documents are available in PDF or MSÔ Word formats from the Western Wood Preservers Institute at

Another wood preservative, copper-8-quinolinolate (solubilized) (Cu-8), is commonly used for above-water components of shorezone structures. Studies of the effects of Cu-8 on aquatic organisms were not located in the course of this review. Cu-8 is considered non-toxic, and is approved for use by the U.S. Food and Drug Administration in wood-preserving applications where treated wood may contact foodstuffs (American Wood-Preservers' Association 1962). It has also been found to not leach from treated textiles in running water (American Wood-Preservers' Association 1962).3. Remaining Issues

Two issues pertaining to piers, docks and bulkheads in the near-shore environment have not yet been addressed. Hydrocarbon-contaminated sediments may be disturbed during new pier, dock or bulkhead construction. There is a lack of published research regarding the location and degree of hydrocarbon sediment contamination levels within Lakes Washington and Sammamish, and a lack of research related to the effects of disturbance of hydrocarbon-contaminated sediments on aquatic organisms.

Household or industrial cleaning and preserving agents that may be applied to piers and docks could have adverse effects on aquatic organisms. No published research on this topic was identified. Additionally, the quantity of chemicals used for this purpose, and the types of chemicals used are unknown. The potential for adverse impacts from household cleaning products is exemplified by the recent fish kill in Thornton Creek, which was suspected to have been caused by a concrete cleaner, treated swimming pool water, or a combination (Birkland 2000).

The additional issue of the potential impacts of lawn-care products on aquatic systems was also not addressed due to time constraints. Many lawn-care products are labeled with warnings of the hazards to aquatic organisms that could result from the inappropriate use of those products. The propensity of the typical waterfront landowner to have a manicured lawn indicates the probability that a variety of pesticides, herbicides, and fertilizers may be routinely applied to those lawns.

Citations we did not collect, but would like to include.

Howard, H.H. and R. Stewart. 1968. Water pollution by outboard motors. Conservationist, June-July, 6-8 and 31.

Hunnefeld, G.R. 1966. Oil pollution in surface waters caused by the operation of outboard motors. Dt. gewasserk. Mitt., 10, 57-59.

Shuster, W. W. and L. Clesceri. 1974. Effects of exhaust from two-cycle outboard engines. U.S. Environmental Protection Agency, Environmental Protection Technology Series, EPA-670/2-74-063.